Transformation/dissolution characteristics of cobalt

and welding fume nanoparticles in physiological and

environmental media: surface interactions and trophic

transfer

Nanxuan Mei

Doctoral Thesis in Chemistry

KTH Royal Institute of Technology

Stockholm 2020

Denna avhandling är skyddad enligt upphovsrättslagen. Alla rättigheter förbehålles.

Copyright © Nanxuan Mei, Stockholm 2020. All rights reserved. No part of this thesis may be

reproduced by any means without permission from the author.

KTH Royal Institute of Technology

School of Engineering Sciences in Chemistry, Biotechnology and Health

Department of Chemistry

Division of Surface and Corrosion Science

Akademisk avhandling som med tillstånd av Kungliga Tekniska Högskolan framlägges till

offentlig granskning for avläggande av teknologie doktorsexamen fredagen den 25

september, 2020 klockan 14:00 i hörsal F3, Kungliga Tekniska Högskolan, Lindstedtsvägen 26,

Stockholm. Avhandlingen presenteras på engelska.

ISBN 978-91-7873-623-2

TRITA-CBH-FOU 2020:37

The following items are printed with permission:

PAPER I: © 2019 ACS Omega

PAPER II: © 2017 Elsevier

PAPER IV: © 2018 Elsevier

Printed by: Universitetsservice US-AB, Sweden 2020

i

Abstract Nanoparticles (NPs) and nanomaterials (NMs) are present everywhere in the environment.

They can form both as an act of nature and during human activities. Various kinds of NPs and

NMs are engineered for different applications in the ongoing development of nanoscience

and technology. Nowadays, concerns have emerged related to potential adverse effects of

NPs on human health and the environment. Knowledge related to effects induced by more

reactive metal NPs is scarce or even missing in some cases. Such information is crucial for

risk assessments. The focus of this doctoral thesis has therefore mainly been placed on

reactive metal NPs: stainless steel welding fume particles, cobalt (Co) NPs, and solution

combustion synthesized (SCS) Co NPs, to investigate their transformation/dissolution

characteristics in environmental and biological media.

Environmental interaction studies were performed in terms of adsorption of biomolecules

and natural organic matter (NOM) onto the surfaces of the NPs and their influence on

dissolution, agglomeration, and size of the NPs in solution. Trophic transfer of Co NPs was

investigated in an aquatic food web.

The Co NPs rapidly agglomerated and sedimented in solution. Co ions were released from

the NPs in both phosphate buffer solution and in freshwater, dissolution processes that were

influenced by the adsorption of biomolecules and NOM. The trophic transfer of Co in the

aquatic food web was shown to be affected by the extent of both agglomeration and

sedimentation. No biomagnification was observed during the trophic transfer, and the

addition of excreted biomolecules had no effect on the transfer.

The dissolution of stainless steel welding fume particles was studied in PBS. The metal

release data could help estimate the risk assessment of stainless steel welding fume particles.

Keywords: nanoparticles, cobalt, stainless steel welding fume particles, biomolecules, metal

release, trophic transfer, adsorption, agglomeration, algae

ii

Sammanfattning Nanopartiklar (NP) och nanomaterial finns överallt i vår omgivning. De produceras genom

både naturliga och mänskliga aktiviteter. Olika typer av NP används inom tillämpningar

såsom kosmetika och läkemedel. Det saknas dock fortfarande kunskaper om effekten av

reaktiva metalliska NP på människors hälsa och miljö. Därför fokuserar denna

doktorsavhandling på reaktiva metallnanopartiklars beteende i miljö- och biologiska media,

till exempel svetspartiklar från rostfritt stål, koboltnanopartiklar och kobolt som producerats

genom lösningsförbränning.

Adsorptionen av biomolekyler och naturligt organisk material i både fosfatbuffert,

saltlösning och ytvatten studerades och dess påverkan på agglomerering och storlek.

Dessutom undersöktes trofisk överföringen av koboltnanopartiklar i en akvatisk näringskedja

besående av alger, zooplankton och fisk. Vidare studerades hur adsorption av utsöndrade

biomolekyler från zooplanktion påverkade överföringen i näringskedjan.

De studerade reaktiva metallnanopartiklarna agglomererade och sedimenterade snabbt i

lösning. Metalljoner frisattes från NP i både fosfatbuffertlösning och ytvatten. Adsorptionen

av biomolekyler och naturligt organiskt material påverkade upplösningen av

metallnanopartiklarna. Överföringen av koboltnanopartiklarna i den akvatiska näringskedjan

påverkades av agglomerering och sedimentation av NP, och det var ingen bioackumulering

under den trofiska överföringen. Tillsatsen av utsöndrade biomolekyler påverkade inte den

trofiska överföringen av koboltnanopartiklar.

Svetsrökpartiklar från rostfritt stål studerades med avseende på upplösning i PBS (simulerad

lungvätska). Studien kan hjälpa till att uppskatta risken av svetsrökpartiklar från rostfritt stål

genom att observera frisättningen av Cr(VI) från partiklarna.

Nyckelord: nanopartiklar, kobolt, rostfritt stål svetsrök partiklar, biomolekyler,

metallfrigöring, trofisk överföring, adsorption, agglomerering, alger

iii

Preface This doctoral thesis aims to understand the transformation of mainly reactive metal NPs

(stainless steel welding fume and Co) in terms of biomolecule/natural organic matter

adsorption, dissolution, and trophic transfer in environmental and biological media.

Relatively inert Co3O4 and WC NPs were studied as comparison. Reactivity is in this case

defined as relatively rapid dissolution rate of the NPs, in general showing significant

dissolution (> ca. 10%) already after 24 h in solution. The research strategy is schematically

illustrated below.

The thesis has been performed within the framework of the Mistra Environmental

Nanosafety program, phase I and II, initiated by the Swedish Foundation for Strategic

Environmental Research (MISTRA). The summary of this thesis aims to reach a large

interdisciplinary audience with different knowledge including e.g. the members in the Mistra

program, experts within the area of NPs and surface chemistry, toxicologists and regulators

interested transformation/dissolution characteristics of NPs and potential adverse effects

and risks, as well as managers in hard metal industries.

iv

List of summarized papers

I. Influence of Biocorona Formation on the Transformation and Dissolution of

Cobalt Nanoparticles under Physiological Conditions

N. Mei, J. Hedberg, I. Odnevall Wallinder, and E. Blomberg

ACS Omega, 2019, 4(26): 21778-21791.

II. Nanoparticles of WC-Co, WC, Co and Cu of relevance for traffic wear

particles – Particle stability and reactivity in synthetic surface water and

influence of humic matter

Y. Hedberg, J. Hedberg, S. Isaksson, N Mei, E. Blomberg, S Wold, I Odnevall

Wallinder

Environmental pollution, 2017, 224: 275-288.

III. Food web transfer of cobalt nanoparticles in algae, zooplankton, and fish

N. Mei, J. Hedberg, M. Ekvall, E. Kelpsiene, L. Hansson, T. Cedervall, E.

Blomberg, I Odnevall Wallinder

To be submitted for publication

IV. Size-separated particle fractions of stainless steel welding fume particles –

a multi-analytical characterization focusing on surface oxide speciation and

release of hexavalent chromium

N. Mei, L. Belleville, Y. Cha, U. Olofsson, I. Odnevall Wallinder, K.-A. Persson, Y.

Hedberg

Journal of Hazardous Materials, 2018, 342: 527-535.

Results from the following manuscript are partly included in the summary of

this thesis:

V. Comparing the reactivity of three different types of cobalt nanoparticles

towards natural organic matter in freshwater

N. Mei, A. Khort, J. Hedberg, T. Chang, E. Blomberg, I. Odnevall Wallinder

Manuscript in preparation

v

Work not included in this thesis

VI. Airborne wear particles generated from conductor rail and collector shoe

contact - influence of sliding velocity and particle size

Y. Cha, Y. Hedberg, N. Mei, U. Olofsson

Tribology Letters, 2016, 64(3): 40.

VII. Mechanical surface smoothing of micron-sized iron powder for improved

silica coating performance as soft magnetic composites

P. Slovenský, P. Kollár, N. Mei, M. Jakubčin, A. Zeleňáková, M. Halama,

I. Odnevall Wallinder, Y. Hedberg

Submitted for publication

VIII. Transformation/dissolution of cobalt nanoparticles in biological media –

relevant for human health and environment

N. Mei, J. Hedberg, E. Blomberg, I Odnevall Wallinder

26-30 May 2019, poster presentation, SETAC Europe 29th Annual Meeting in

Helsinki, Finland.

IX. Minimized risk for exposure and release of harmful substances when

welding stainless steels

Z. Wei, S. McCarrick, V. Romanovski, J. Theodore, N. Mei, K.-A. Persson,

O. Runnerstam, H.L. Karlsson, I. Odnevall Wallinder, Y. Hedberg

23-25 October 2019, poster presentation, Materials and Formulations at

Biointerfaces, a symposium on surface chemistry and materials science, Malmö,

Sweden.

X. Importance of electrochemical and surface characteristics of a range of

metal nanoparticles for environmental fate

J. Hedberg, Y. Hedberg, N. Mei, E. Blomberg, I. Odnevall Wallinder

5-9 November 2018, Nanosafe, Grenoble, France.

XI. Bioaccessibility testing and characterization of five molybdenum

compounds

Z. Wei, N. Mei, X. Wang, J. Hedberg, I. Odnevall Wallinder, Y. Hedberg

Technical final report, commissioned by the International Molybdenum

Development Association, December 2019.

vi

Author’s contributions to the appended papers

Paper I: Experimental work: ATR-FTIR, AAS, particle digestion, PCCS, Zeta potential. Major

part of planning and design of experimental set-up, data evaluation/ interpretation and

preparation of the manuscript.

Paper II: Experimental work on dihydroxy benzoic acid (DHBA) adsorption using ATR-FTIR.

ATR-FTIR part of writing of the manuscript.

Paper III: Experimental work: ATR-FTIR, particle exposure, quantification of Co uptake by

Daphnia magna (digestion and AAS) and part of fish organ samples digestion and AAS

quantification, PCCS, Zeta potential. Major part of planning and design of experimental set-

up, data evaluation/ interpretation and preparation of the manuscript.

Paper IV: Part of experimental work including CV, particle exposure, particle digestion and

AAS analysis. Minor part of planning and design of experimental set-up, part of data analysis

and preparation of the manuscript.

Results from the following manuscript are partly included in the summary of this thesis:

Paper V: Part of experimental work including particle exposure, ATR-FTIR, AAS, particle

digestion. Part of planning and design of experimental set-up, data evaluation/

interpretation and preparation of the manuscript.

vii

Acknowledgements

I am happy to thank all of you that helped me during my PhD studies. Without your

encouragement and support, I could not have accomplished my PhD.

Firstly, I would like to express my greatest gratitude to my main supervisor Assoc. Prof. Eva

Blomberg for your support and patience. I met lots of difficulties during the PhD study, and

you always believed in me and helped me as much as you could. You are very selfless,

intelligent and kindhearted both in science and in life. I am very proud that I had such an

excellent supervisor during my PhD study. I wish that you will always be happy, peaceful and

never feel lonely. I will never forget the 4 years that we have spent together, and I hope that

I can collaborate with you again.

I would like to express my sincere gratitude to my co-supervisor Prof. Inger Odnevall

Wallinder. You could always provide wonderful ideas when I met problems during the PhD

project, and helped me to control the right direction for my study. You have profound

knowledge, especially for corrosion which helped me to solve lots of questions. It was very

lucky for me to be your PhD student since you let me understand how scientific research

goes and how to find the key point of a project.

I would also appreciate my co-supervisor Dr. Jonas Hedberg. We spend most of the time

together during my PhD study, from the design of experiments to analyzing the data. Every

time I got problems or troubles for experiments, you were always there and gave me

support as much as you could. You are a very hard-working researcher, and I hope you will

achieve success in the future. I will be very happy if we can collaborate again.

I also want to say thanks to Dr. Yolanda Hedberg. You are the one who brought me to the

wonderful division of Surface and Corrosion Science. I still remembered when we met in the

corrosion course for master students, and you offered me the opportunity to work with you

on the welding fume particles project. It was the first time for me to get in touch with

scientific research and collaborating with you made a good impression on me about science,

which firmed my mind to continue my PhD study. I am glad that you were my supervisor for

the master thesis.

Thanks to Prof. Per Claesson. You are a great expert on surface chemistry and a successful

teacher. I like your course, and you made me see that surface chemistry is interesting

instead of only boring concepts and formulas.

Thanks to Prof. Mark Rutland. You were the teacher of my first PhD study course. You are

very humorous, and the lectures were very impressive. I appreciate the basic surface

chemistry knowledge that I learnt from you.

Thanks to Prof. Christofer Leygraf. You gave a wonderful course on corrosion science which

helped me a lot during my PhD study.

Thanks to Assoc. Prof. Eric Tyrode. I got the knowledge of IR and Raman from you. The

advanced surface chemistry course was useful, as well. You are very intelligent and selfless.

viii

You always helped me when I got questions, even though you were busy with your own

business.

Thanks to Dr. Gunilla Herting. You built very strict rules in the lab, which are very necessary.

We could not have such a nice structured lab without your contribution. You helped me a lot

with AAS analysis. I feel lucky that I worked in the lab together with you.

Thanks to Zheng Wei, Xuying Wang, Aliaksandr Khort, Amanda Kessler and Tingru Chang. We

worked together in Inger’s group, and you made the working environment nice and friendly.

We always helped each other; I will have a good memory to be a colleague with you all.

Thanks to Gen Li, Zheng Wei and Yonggang Yang. You were great roommates, and we had a

happy life in Kungshamra 3 in Solna.

Thanks to all former and present colleagues at Division of Surface and Corrosion for a

friendly and wonderful working environment. I will miss our Tuesday seminars, Friday “fika”

and team-building activities.

Warm thanks to my friends in both China and Sweden, Delong Zhao, Tianbo Xu, Qingxin

Zhang, Qizheng Zhang, Juyang Sun, Nan Zhu, Heran Jiang, Huifeng Huang, Simin An and Tijie

Xu for your encouragements and all the nice moments we have shared.

Thanks to the colleagues in the Mistra Environmental Nanosafety project for nice

cooperation.

The Chinese Scholarship Council (CSC) is gratefully acknowledged for the financial support

for my PhD study.

Last but not least, I want to thank my father, Qi Mei, my mother Man Zhao and my girlfriend

Jie Cheng. You gave me great support and love that encouraged me all the time. I hope you

all will be happy and healthy.

最终我要感谢我的父亲，梅琪，母亲，赵满，以及我的女朋友程洁。没有你们的爱和

支持我很难走到如今的地步，感谢你们陪伴我成长。希望你们永远健康幸福。另外要

特别感谢我的姥爷，赵庭耀，我的姥姥，李秀琴，感谢你们陪伴我成长，给予我一个

温暖幸福的家庭。

ix

Abbreviations AAS Atomic absorption spectroscopy

ATR-FTIR Attenuated total reflection Fourier transform infrared

spectroscopy

BAF Bioaccumulation factor

BET Brunauer–Emmett–Teller surface area measurement

BMF Biomagnification factor

BSA Bovine serum albumin

BSE Backscattered electrons

Co Cobalt

Cr Chromium

Cu Copper

CV Cyclic voltammetry

2,3-DHBA 2,3-dihydroxybenzoic acid

3,4-DHBA 3,4-dihydroxybenzoic acid

DLVO Derjaguin-Landau-Verwey-Overbeak theory

EDL Electrostatic double-layer force

EDS Energy dispersive spectroscopy

Fe Iron

FW Freshwater

Mn Manganese

MQ MilliQ, Ultrapure water

Ni Nickel

NMs Nanomaterials

NOM Natural organic matter

NPs Nanoparticles

NTA Nanoparticle tracking analysis

OCP Open circuit potential

PBS Phosphate buffered saline

PCCS Photon cross correlation spectroscopy

x

PIGE Paraffin-impregnated graphite electrode

SE Secondary electrons

SEM Scanning electron microscopy

SCS Solution combustion synthesis

SGF Simulated gastric fluid

TEM Transmission electron microscopy

TOC Total organic carbon

TW Tap water

vDW van der Waals force

WC Tungsten carbide

WC-Co Tungsten carbide with cobalt

XDLVO Extended DLVO theory

XPS X-ray photoelectron spectroscopy

XRD X-ray diffraction

xi

Table of contents

Abstract .................................................................................................................................................... i

Sammanfattning .......................................................................................................................................ii

Preface ..................................................................................................................................................... iii

List of summarized papers ...................................................................................................................... iv

Work not included in this thesis ............................................................................................................... v

Author’s contributions to the appended papers ..................................................................................... vi

Acknowledgements ................................................................................................................................ vii

Abbreviations .......................................................................................................................................... ix

Table of contents ..................................................................................................................................... xi

1. Motivation and scope .......................................................................................................................... 1

2. Introduction ......................................................................................................................................... 4

The relevance to the United Nations Sustainable Development Goals .............................................. 4

Origin of nanoparticles ........................................................................................................................ 4

Comparison between NPs and massive (bulk) materials .................................................................... 5

Transformation of nanoparticles in different fluids ............................................................................ 7

Nanoparticle agglomeration............................................................................................................ 7

Interactions with organic matter – change in surface properties ................................................. 10

Dissolution of the NPs – metal release ......................................................................................... 11

Nanoparticle biouptake and trophic transfer ............................................................................... 12

Adverse effects and toxicity of nanoparticles ................................................................................... 14

Which parameters determine the hazard of NPs? ............................................................................ 15

The properties of the NPs.............................................................................................................. 15

Metal release from the metal NPs surface .................................................................................... 16

Information on the metal NPs used in this PhD-project ................................................................... 16

3. Experiments and Techniques ............................................................................................................ 19

Speciation modelling of released Co in solution ............................................................................... 21

Attenuated Total Reflection Fourier Transform Infrared Spectroscopy (ATR-FTIR) ......................... 21

Atomic absorption spectroscopy (AAS) – dissolution / metal release .............................................. 23

Exposure of nanoparticles in solution for metal release determinations ......................................... 24

X-ray Photoelectron Spectroscopy (XPS) ........................................................................................... 25

Electrochemistry – Open Circuit Potential (OCP) and Cyclic Voltammetry (CV) ............................... 25

Transmission Electron Microscope - Energy Dispersive Spectrometer (TEM-EDS) ........................... 26

Scanning Electron Microscopy (SEM) ................................................................................................ 27

xii

X-ray diffraction (XRD) ....................................................................................................................... 27

Photon Cross Correlation Spectroscopy (PCCS) ................................................................................ 28

Nanoparticle Tracking Analysis (NTA) ............................................................................................... 28

Zeta potential .................................................................................................................................... 28

Brunauer–Emmett–Teller (BET) – surface area measurement ......................................................... 29

4. Key Results and Discussion ................................................................................................................ 30

4.1. Characterization of particles before exposure. (Papers I-V) ...................................................... 30

4.2. Fate and transformation of Co NPs under simulated physiological conditions: influence of

biocorona formation (Paper I). .......................................................................................................... 33

4.3. Fate and transformation of metal NPs in simulated environmental conditions: influence of

ecocorona formation (Paper II and unpublished results) ................................................................. 39

Nanoparticles of relevance in traffic settings ................................................................................ 39

Core-shell, oxide, and composite Co nanoparticles behave differently in freshwater in terms of

dissolution and NOM adsorption. ................................................................................................. 40

4.4. Trophic transfer of Co NPs in the aquatic food web (Paper III) ................................................. 47

4.5. Dissolution of stainless steel welding fume particles in PBS solution (Paper IV) ....................... 55

Concluding remarks ............................................................................................................................... 59

Future work ........................................................................................................................................... 61

References ............................................................................................................................................. 63

1

1. Motivation and scope Nanomaterials (NMs) are ubiquitous and present in virtually every breath, every drink, and

every bite.[1] There are both natural nanoparticles (NPs) and anthropogenic NPs, for

example in fumes from volcanic eruptions and as a result of coal combustion. NPs are

defined as having at least one dimension smaller than 100 nm.[2]

Engineered NPs and NMs are produced to improve specific properties of products in various

applications ranging from human health to electronics, materials, energy, water, and food

production.[1] One of the main advantages of using NMs is the increased reactivity per mass

of materials which allows faster processes, and, in many cases, smaller amounts of materials

needed compared to conventional solutions.[3] NPs can also have different physico-chemical

properties compared with bulk material, such as color, solubility, mechanical strength,

optical properties, diffusivity, crystallinity, etc.[4] Surface atoms are not as stable as bulk

atoms, and the large fraction of surface atoms leads to some specific properties such as

higher dissolution rate compared with the equivalent bulk material.[5]

Studies of effects of NPs on human health and environment have become an active research

area due to for example air pollution issues and increased production of engineered NPs.

Thus, an increasing use of NPs creates new challenges for regulation and testing. The

physico-chemical properties of the NPs will influence their toxicity; for instance, the toxicity

of silver NPs is affected by the free silver (Ag) ion concentration in solution.[6] In order to

estimate Ag NP toxicity, it is hence important to determine whether the given chemical

setting enhances or reduces the dissolution and availability of free silver ions in solution.[7-9]

Since NPs are very sensitive to the surrounding chemical environment, changes in for

example solution pH, ionic strength, and composition will influence the reactivity of the

NPs.[9] It is therefore very important to understand the influence of the surrounding media

when evaluating environmental or health risks induced by the exposure to NPs. For instance,

knowledge on transformations of NPs (e.g. dissolution, heteroagglomeration) for a given

exposure route provides useful information on their environmental fate. NPs, released ions

and formed complexes can be taken up by organisms,[10] and their biouptake is affected by

their surface characteristics and reactivity that in turn are related to the chemistry of the

surrounding environment.

A main focus of this PhD thesis has been to generate an in-depth fundamental

understanding of interactions between metallic NPs and biomolecules and natural organic

matter. Dissolution and adsorption mechanisms were studied in terms of their importance

for biomolecule-induced corrosion and metal release of technically relevant metallic NPs

(mainly cobalt) in simulated physiological, biological and freshwater solutions. A main focus

was to elucidate processes and mechanisms of reactivity, biomolecule interactions and

metal release/dissolution characteristics, and relate these aspects to possible toxic effects.

The thesis focuses on understanding the behavior of different metal and metal oxide NPs

(stainless steel welding fume particles, Co NPs, Co3O4 NPs, Co SCS NPs, WC NPs, WC-Co NPs,

Cu NPs) at different conditions of relevance for environmental and human exposure

conditions, schematically illustrated in Figure 1. The transformation of Co NPs is an example

2

of reactive metal NPs that up to now has not been as widely studied compared to for

example Ag or Au NPs. Different types of Co NPs were investigated, e.g. Co NPs with a

surface oxide (bulk/shell), Co3O4 NPs, and Co NPs with a carbon layer (SCS). Stainless steel

welding fume NPs were studied to understand the release of Cr(VI) from the NPs and the

surface speciation. The ultimate goal was that findings of this thesis can contribute with

knowledge that can be used when performing risk assessments of metal-containing NPs.

The thesis can be divided into four main parts:

Interactions between Co NPs and biomolecules in solution to assess how these

interactions affect their dissolution and agglomeration characteristics.

Effects of NOM on the transformation of NPs of Co, Co3O4, and Co with a carbon layer

(SCS Co) in terms of changes in surface characteristics and dissolution in solution.

Parallel studies on WC-Co NPs (relevant to NP emissions in traffic) compared WC NPs

and Cu NPs in terms of transformation in the presence of small-sized NOM.

Transfer and biomagnification of Co NPs through an aquatic food web, including

algae (Scendesmus sp.), zooplankton (Daphnia magna) and fish (Crucian Carp) and

influence of the interaction between biomolecules excreted by Daphnia magna and

Co NPs on the transformation of Co/Co NPs in trophic levels.

Effects of welding settings on the surface speciation and dissolution of welding fumes

NPs in phosphate buffered saline.

Figure 1. Schematic summary of the different research projects included in the doctoral thesis.

A systematic approach combining surface chemistry, corrosion and solution chemistry with a

multitude of state-of-art surface analyses was employed in the research projects. A range of

3

instrumental techniques was used including XPS (X-ray photoelectron spectroscopy), XRD (X-

ray diffraction), BET (Brunauer-Emmett-Teller surface area analysis), TEM (Transmission

electron microscopy) and SEM (Scanning electron microscopy) was used to characterize the

NPs. ATR-FTIR (Attenuated total reflection-Fourier transform infrared spectroscopy) was

employed to investigate the adsorption of biomolecules and NOM onto the NPs. AAS

(Atomic absorption spectroscopy) was used to determine total concentrations of released

metal ions related to NP dissolution. PCCS (Photon cross correlation spectroscopy) and NTA

(Nanoparticle tracking analysis) were used to determine the NP size distribution. Zeta

potential measurements were conducted to provide information on the apparent surface

potential, which is related to surface charge.

4

2. Introduction The relevance to the United Nations Sustainable Development Goals This PhD-project is focused on the transformation of metal-containing NPs at aqueous

environmental conditions and at simulated physiological conditions in terms of

biomolecule/natural organic matter adsorption and dissolution. Improved knowledge on the

environmental fate and importance of chemical setting conditions on the

transformation/dissolution of metal-containing NPs are crucial aspects to consider for e.g.

accurate risk predictions and for legislative actions. For example, it was found that the

welding fume NPs can release Cr(VI) in simulated blood serum solutions, which can result in

adverse effects on human health. Dispersion of e.g. Co NPs from car studs generated via

different wear processes can both induce negative effects on humans if inhaled as well as be

toxic to aquatic organisms at environmental settings. Findings in this PhD thesis contribute

to the sustainable development goals of “Good Health and Well-Being, no. 3”, “Clean Water

and Sanitation, no. 6” and “Life Below Water, no. 14”. For instance, the studied NPs could

potentially be included in the hazardous chemicals in target 3.9 “By 2030, substantially

reduce the number of deaths and illnesses from hazardous chemicals and air, water and soil

pollution and contamination”. Since the NPs behavior studied in this thesis are all in aquatic

condition, the results could be contribute to target 6.3 “By 2030, improve water quality by

reducing pollution, eliminating dumping and minimizing release of hazardous chemicals and

materials, halving the proportion of untreated wastewater and substantially increasing

recycling and safe reuse globally” by identifying patterns in environmental fate NPs which

can be used for fate modelling and risk assessment. Furthermore, the biouptake and trophic

transfer of Co are investigated in the thesis, which is related to target 14.1 “By 2025, prevent

and significantly reduce marine pollution of all kinds, in particular from land-based activities,

including marine debris and nutrient pollution” by providing information on possible

biomagnification of Co upon dispersion of Co NPs.

Origin of nanoparticles

NPs are present everywhere in the environment, and their origin can be divided into three

main categories: natural, incidental, and engineered.

Several processes in nature generate particles and NPs. 90% of the NPs in the atmosphere

has a natural origin.[11] A dust storm is the largest single source for natural NPs.[12] Half of

the dust in terms of mass consists of particles smaller than 2.5 µm, and the smallest particle

size can be less than 100 nm.[11, 13] A large number of NPs is formed as a result of forest

fires, burning of trees and grass,[14] and generated during volcanic eruptions.[11]

Human activities can result in incidentally generated NPs, for example during welding,

mechanical processes, and combustion.[12] Vehicles in traffic settings are a main source of

NPs in urban areas. Diesel engines typically generate NPs in size range of 20-130 nm, and

gasoline engines in the range 20-60 nm.[15] Carbon nanotubes have also been observed to

form during diesel- and gas combustion.[16, 17] NPs can also be generated at indoor

conditions and be caused by human activities such as cooking.[18, 19] Cigarette smoke is

5

another origin of NPs that typically are sized between 10 nm and 700 nm, with a mean value

of 150 nm.[20]

Certain NPs are engineered for specific applications. Such particles have been used for a long

time, for example in cosmetics in ancient Egypt, though without the knowledge that they

actually were nanosized.[12] Examples of engineered NPs include TiO2 NPs, used in white

pigments, food colorants, sunscreens and cosmetic creams,[21] and Ag NPs used for

antibacterial applications.[22] Coatings containing NPs are used for many applications, for

example TiO2 NPs coated onto stainless steel to hinder corrosion, and different TiO2 NP-

containing coatings are used in drug delivery.[23-25]. More and more nanostructured

materials are used in products on the market, including super hydrophobic coatings,

catalysts, and biomaterials. TiO2 is one of the most widely used engineered NPs with a global

production number of 3000 tonnes/year (median value) reported in 2012. Corresponding

numbers were 5500 tons/year for SiO2, 550 tonnes/year for ZnO and 300 tonnes/year for

carbon nanotubes. Approximately 70% of the total amount of TiO2 and ZnO NPs were used

for cosmetic applications. Carbon nanotubes were mainly used for material production and

in batteries.[26]

Comparison between NPs and massive (bulk) materials

Nano-specific properties can broadly be divided into surface effects and quantum effects.[2]

Atoms at the surface have different properties from atoms in the bulk due to different

interaction energies between the atoms, Figure 2. The surface atoms have fewer neighbors

than bulk atoms, which results in an excess in energy at the surface compared to the bulk.[2]

The consequence of having fewer neighbor atoms is that atoms at the surface in general are

not as stable as bulk atoms. Due to their small size, NPs have a larger fraction of surface

atoms compared with bulk materials. The ratio of surface area to particle volume can

estimate the fraction of surface atoms. As an example, for a 60 nm sized particle, the ratio is

1000 times higher compared to a particle sized 60 µm.[12] The number of surface atoms of

NPs is hence large, which makes NPs more reactive than the corresponding massive material.

The NPs are thus more prone to oxidize, which usually results in a higher dissolution rates

compared to the corresponding micro-meter sized particles.

6

Figure 2. A schematic illustration of the difference between surface- and bulk atoms for a

bulk material (left) and a NP (right). The red arrows correspond to the interaction energy

between atoms.

The dissolution rate of particles is also related to the particle size according to the so-called

Kelvin effect.[27] This effect predicts that smaller particle sizes result in higher solubility and

dissolution rates compared to the bulk material due to increased surface curvatures of the

smaller particles.

The corrosion potential of metal NPs depends not only on the work function of the massive

metal but also on the size and charge of the NP, the dielectrics of the solvent and on the

characteristics of adsorbed molecules. Metal NPs in solution are often synthesized with

anions adsorbed to the surface, conditions that in many cases influence both the corrosion

potential and the apparent surface charge.[28]

Due to their small size, NPs have a high specific surface area and higher surface energy from

which follows an increased tendency for adsorption of inorganic ions, organic molecules,

and/or water molecules from solution in order to reduce surface energy.[27]

The small size of NPs can also influence their crystal structure.[27] For example, the

thermodynamically stable phase for TiO2 is rutile for particle diameters exceeding 14 nm,

whereas anatase is the stable phase for particle sizes less than 14 nm.[29]

The quantum effect has not been in focus within the framework of this thesis. This effect is

related to that NPs show a discontinuous behavior due to quantum confinement effects with

delocalized electrons. An example is quantum dots which consist of synthesized

nanostructures with particle sizes as small as only a few nm.[30] Another example of the

quantum effect is the appearance of magnetic moments in NPs that are non-magnetic at

bulk conditions, such as NPs of gold (Au), platinum (Pt) or palladium (Pd).[2]

7

Transformation of nanoparticles in different fluids The transformations of NPs in environmental and biologically relevant fluids must be

considered to enable relevant predictions of effects and risks induced by the

dispersion/exposure of NPs in the environment and on humans (Figure 3). This thesis focus

mainly on the transformation behavior of different metal-containing NPs including changes

in surface characteristics, agglomeration behavior and dissolution in solution, and possible

trophic transfer in the food web.

Figure 3. Schematic illustration of selected transformation processes of NPs that influence

their fate, transport, and environmental toxicity.

Nanoparticle agglomeration

Depending on the aquatic exposure setting, it is likely that NPs will agglomerate in different

ways and not really exist as individual (pristine) particles. These processes are caused by the

interaction between NPs (homoagglomeration between similar particles) as well as between

NPs and constituents in the surrounding fluids (heteroagglomeration). The classical DLVO

theory (Derjaguin-Landau-Verwey-Overbeak) from colloidal science may be used as a first

approach to predict the agglomeration of NPs, as shown in Figure 4.[31] In the DLVO theory,

the attractive van der Waals force (vdW) and the repulsive electrostatic double-layer force

(EDL) are assumed to be additive. The DLVO force is obtained by summation of these to

forces. Between similar NPs (homoagglomeration), the following information can be drawn

from the DLVO theory:

The vdW force is always attractive in any media, and it originates from the

interaction between induced and/or permanent dipoles. It depends on the particle

8

size and the material properties (i.e. dielectric constant and refractive index), which is

expressed by the Hamaker constant. When considering metal NPs, the attractive

vdW force will be substantial in solution due to a high Hamaker constant. This

originates from conductive and polarizable metal NPs, which results in high dielectric

constants and refractive indices and therefore a high Hamaker constant. Metal NPs in

an aqueous solution without any surface modification will hence always agglomerate

and subsequently sediment due to gravitational forces.[32]

The EDL force is always repulsive and originates from the overlap of the ion cloud

outside the charged NP surface, which increases the osmotic pressure and that

results in a repulsion. In aqueous solutions, the NPs are always charged due to

dissociation of the surface groups or adsorption of ionic species from the solution.

Most metal NPs spontaneously form surface oxides in contact with the ambient air

and aqueous solutions, forming a hydroxylated surface that results in a charged

surface.[33]

Due to the high Hamaker constant for metal NPs, the vdW force will dominate over the EDL

force. This results in a rapid agglomeration and subsequent sedimentation of the NPs in

solution, effects that have been shown in several studies investigating different kinds of

metal NPs.[34-37]

Although the DLVO theory successfully has been used for explaining the colloidal stability for

micron-sized particles, the validity of this theory for NPs is not evident. The DLVO theory

may not be suitable for very small particles since the surface curvature is too substantial to

assume a flat surface. The chemical composition will furthermore affect both the Hamaker

constant and the surface charge, which influences the magnitude of the vdW force and the

EDL force, respectively. The crystal structure as well as the shape of the particles can also

influence the Hamaker constant and the surface charge. Rectangular rods, and cylinders

have for example been reported to have larger attractions force than spherical particles.[32]

9

Figure 4. Schematic drawing of the DLVO interaction.

In a complex reality it may however be difficult to use the classical DLVO theory to predict

agglomeration of NPs since many factors need to be taken into account. As an example,

some engineered NPs are produced with organic coatings adsorbed to the surface. Its

presence may, at least for a certain time period, due to steric forces prevent from particle

agglomeration. Steric forces originate from volume restrictions and inter-penetration effects

of the adsorbed layer that result in a repulsion. Since environmentally dispersed NPs always

are in contact with natural organic matter (NOM), any adsorption of NOM to the NP surfaces

can give rise to steric repulsions. The adsorption of NOM has also been reported to make the

agglomeration reversible.[38] Adsorption of organic molecules lowers the attractive vdW

force since the Hamaker constant for the organic layer is significantly lower compared to the

metal NPs. The classical DLVO theory is hence not sufficient to predict the stability of NPs in

solution. The extended DLVO theory (XDLVO) may be used instead as this theory also

considers additional short-range forces, such as bridging, osmotic, steric, hydrophobic Lewis

acid-base, and magnetic forces.[32].

When considering the agglomeration of particles in the environment, it is also important to

understand effects of the chemical setting, such as pH and ionic strength. Surface charge

titration and EDL screening are two primary effects in which pH and ionic solutes may

promote NP agglomeration. For each particular system, the pH at which the H+ and OH−

concentrations cause suspended particles to obtain a neutral charge is called the point of

zero charge (PZC). As the pH moves toward this point, the EDL repulsion decreases, and

particle agglomeration is promoted by vdW attraction. A high ionic concentration decreases

the Debye length, which results in a reduced range of the EDL repulsion.[32]

10

It is hence very difficult to predict NPs agglomeration in solution using common rules.

Nevertheless, it is an important behavior that cannot be ignored when investigating the

environmental fate and behavior of dispersed NPs to different environmental settings.

Interactions with organic matter – change in surface properties

Prevailing environmental conditions need to be addressed when considering the surface

characteristics of metal-containing NPs since surface interactions to different extent will take

place with various ligands and biomolecules and which influence the environmental fate and

toxic potency of such NPs.

In order to understand biomolecule and NOM adsorption, the first step is to know why

adsorption occurs. Several intermolecular interactions can influence the adsorption of

biomolecules on surfaces including ionic (electrostatic) interactions (both repulsive and

attractive), hydrogen bonding, hydrophobic interaction, and van der Waals forces.[39]

Ionic interactions: These are also known as coulomb interactions. It is an effective

contribution when the sorbent surface and adsorbing molecule are electrically

charged. In aqueous media, biomolecules and NPs are usually charged. The coulomb

interaction can be either attractive or repulsive.

Hydrogen bonding: H-bonding is effective at surfaces with the presence of electron-

donating or electron-accepting groups. The H-bonding contribution in aqueous media

for adsorption is a critical condition due to the water H-bonding.

Hydrophobic interactions: Hydrophobic interaction occurs in aqueous media. It can

simply be described by that surfaces in water prefers to be hydrophilic to make the

free energy of the system lower. Since biomolecules contain several hydrophilic and

hydrophobic functional groups, their adsorption onto the NPs surface can make the

system more thermodynamically stable.

Van der Waals forces: van der Waals force is a general force to describe attraction

intermolecular forces between molecules. Dispersion (London-vdW) interactions are

always operating. Dipolar interactions (Debye-and Keesome-vdW) acting between

polar and polarizable components.

The driving force of biomolecule adsorption is widely known, and the adsorption of

biomolecules is a process that includes several steps.

1. Transport of the biomolecule towards the NP surface.

2. Attachment of the biomolecule onto the surface. The biomolecules have several

different functional groups, and usually, the molecules have a spatial orientation in

the solution. Adsorption onto hydrophobic surfaces will occur with hydrophobic

groups and onto hydrophilic surfaces with hydrophilic groups.[40-42]

3. Biomolecules may change conformation and re-orient upon adsorption onto surfaces.

The change of conformation needs time and will result in irreversible adsorption. The

conformational changes are affected by the surface coverage of the biomolecules

and the structural stability of the biomolecule.[43-45]

11

4. The detachment of the biomolecule from the surface. The adsorbed biomolecule on

the nanoparticle can be detached due to structural changes or ligand-induced

processes forming metal-ligand (biomolecule) complex.[46, 47]

Biomolecule adsorption generally changes the surface characteristics and reactivity of the

NPs and hence the fate and toxic potency of NPs in the environment, effects for instance

reported for metal NP-NOM interactions.[48] It is hence very important to consider surface

interactions with different ligands such as biomolecules at the given exposure setting when

assessing environmental risks related to the dispersion of metal NPs.[36, 49, 50]

Dissolution of the NPs – metal release

As schematically illustrated in Figure 5, metal dissolution can be divided into two main types:

electrochemical dissolution where ions leave the metal surface (typically oxidized) due to a

corrosion process, and chemical dissolution due to for example ligand-induced processes in

which metal ions form a complex with adsorbed molecules or ions that desorb without

electron transfer.

Corrosion is the driving force for a metal to reach its most thermodynamically favorable

state. The oxidation of the metal is accompanied by a reduction reaction, where the oxygen

evolution is the most important reaction at natural pHs in aerated solutions. If the corrosion

product (in general a metal oxide) has a high solubility in solution, dissolution will continue

until saturation of soluble metal ions in solution is obtained for the given exposure

conditions. However, poorly soluble corrosion products and protective surface oxides can

hinder the corrosion process.

Figure 5. A schematic illustration of corrosion and metal release processes that can take

place on metal NPs.

Different from corrosion, NPs dissolution processes can involve no electron transfer.

Dissolution of substance A can be described as aA(s) ⇌ bB(aq) + cC(aq), and Ksp (equilibrium

constant of the solubility products):

𝐾𝑠𝑝 = {𝐵}𝑏{𝐶}𝑐

12

where {B} and {C} are the activities of ions in solution. Usually, metal and metal oxide NPs

have very small solubility products and do not follow this process. However, the NPs size can

influence the dissolution, and the Kelvin equation is modified to describe the relationship

between the solubility of spherical particles and the radius. It predicts that the solubility

increases when the radius decreases.

The extent of NP dissolution is also influenced by the presence and concentration of ligands

and molecules in solution. When dissolved metal ions form a labile (easily dissociated) or

strongly bonded complex with e.g. a biomolecule in solution, the equilibrium of dissolution

will be pushed further by increasing the metal solubility, and more metal ions will be

dissolved from the NPs. As previously mentioned, adsorption of biomolecules/organic

matter onto the NPs surface may also influence the extent of dissolution.[51]

However, as metal dissolution typically is governed by a combination of differently induced

chemical- (e.g. proton- and ligand induced) and electrochemical processes, metal dissolution

must be measured and determined for the given exposure setting and cannot be

theoretically calculated based on available solubility data for a similar oxide or compound as

present in the surface oxide of the NPs.

Particle agglomeration will affect the dissolution since it reduces the specific surface area

and increases the radius (the radius of agglomerate).[27]

Nanoparticle biouptake and trophic transfer

The organisms in a food chain are classified into different trophic levels based on their

feeding behavior. An understanding of trophic transfer of NPs would improve both

environmental and health risk assessments. The trophic transfer is described as the

movement of toxicants up through the food web via ingestion of preys by predators. It has

been widely recognized and remains a much-studied ecotoxicological issue.[10] The trophic

transfer of NPs can be described as how the NPs travels inside the food web from the lower

level (e.g. algae) to a higher trophic level (e.g. fish), schematically depicted in Figure 3.[10] It

is necessary to understand how NPs are taken up by different organisms (biouptake) in order

to assess the trophic transfer of NPs, Not surprisingly, the NPs can interact with aquatic

organisms.[52] This type of attachment is related to surface interactions, which depend on

the physical and chemical characteristics of the NPs.

Biouptake is the process where metal NPs or ions/complexes are taken up and remain in an

organism or a cell. Within the context of this thesis, this process does not mean weak

interactions where the NPs can be easily removed or rinsed off (illustrated in Figure 6) from

the organism or cell. The process of biouptake can be divided into three main steps:

i) Diffusion – the metal NP or ion/complex move from the bulk solution to the vicinity

of the surface of the organisms. During the diffusion step, the metal ions might form

complexes with other molecules.

13

ii) Adsorption – when the metal NP or ion/complex come into contact with the

biological interface, they may adhere (heteroagglomerate) to specific sites on the

biological surface.

iii) Internalization –the metal NP or ion/complex can transfer through the cell

membrane and enter the cell.

Figure 6. Schematic illustration of possible biouptake of metal NPs and metal ions/complex in

a cell.

Another concept that is related to biouptake is bioavailability. Bioavailability is the extent to

which a bioaccessible substance (the total amount of in this case metal NPs or

ions/complexes that may be available for uptake) actually can be taken up by a living

organism and cause adverse physiological or toxicological effects. The bioavailability is

related to the chemical form of the substance, in this case if the metal NPs exist as NPs, free

ions and/or form labile or strong complexes with inorganic and organic ligands (chemical

speciation) that can be taken up by the organism. This, in turns, influences the toxic

potency.[53]

If metal NPs are dispersed into an aqueous environment they typically encounter NOM that

in most cases adsorb to different extent onto the NP surface.[48, 49] Adsorbed NOM

molecules can stabilize the NPs in solution and prevent them from agglomerating. These

stabilized NPs can then possibly be directly taken up by cells due to endocytosis.[48] Some

NPs can penetrate the cell membrane directly without any specific receptors due to passive

diffusion. For instance, TiO2 and polystyrene could penetrate the plasma membrane in this

way.[54, 55]

Moving back to the trophic transfer, it is from a risk perspective essential to determine

whether metal ions/complexes or metal NPs can be taken up by an organism, and if they will

be transferred to a higher trophic level. Biomagnification and bioaccumulation are terms

that are used to describe this issue. Biomagnification means that in a higher trophic level,

the amount of hazardous or toxic materials is accumulated in the organism compared to

lower trophic levels in the food chain. In this PhD thesis, the unit g metal/ g dry body weight

has been used to calculate the biouptake of metals in an organism in different trophic levels.

The extent of biouptake can indicate if biomagnification will take place or not. The

14

bioaccumulation factor (BAF) is the ratio of the total metal concentration in an organism

originating from all exposure pathways (including water, sediment, and dietary pathways)

per-unit fresh tissue weight basis compared with the background concentration for the

specific setting.[56] Bioaccumulation of non-essential metals may pose adverse risks on both

humans and other organisms and thereby an essential aspect to consider in risk assessment.

The trophic transfer of Co NPs was in this thesis investigated for an aquatic food web with

algae (Scenedesmus sp.), zooplankton (Daphnia magna), and fish (Crucian carp). Daphnia

magna is commonly used in laboratory investigations to study bioaccumulation of metal NPs

in environmental- and food-relevant settings.[10] The zooplankton continuously ingest

material suspended in the water column and can filter 0.1-1.5 µm sized particles, which

means that ingestion of agglomerated NPs is possible.[58] The trophic transfer of metal NPs

is influenced by ingestion and depuration by the Daphnia magna, as elucidated by earlier

studies on Au NPs. Their uptake of Au NPs was the same regardless of the initial surface

chemistry of the Au NPs, and if the Daphnia were allowed to have a depuration period.[52,

59] Other studies show that TiO2 NPs can be transferred from Daphnia to zebrafish via food

ingestion, though no biomagnification was observed (BMF < 1). [60] BMF was calculated by

ratio of the TiO2 NP concentration in zebrafish (mg kg-1) to that in its diet of Daphnia magna

at steady state.

Adverse effects and toxicity of nanoparticles This PhD thesis focuses on the transformation of metal NPs in different environments. The

obtained results can aid in estimating NPs toxicity and risks. Potential environmental and

health risks related to NPs need to be addressed for exposure scenarios of relevance for e.g.

inhalation, including air pollution, [61-63] since inhalation of NPs will mainly affect the

lungs.[64, 65].

NPs in airborne dust can cause asthma and emphysema.[11] It has also been reported that

the medical visits for respiratory illness increased by more than 50% during the weeks of the

forest fires during the large US wildfire in Humboldt County, California in 1999. On a long-

term perspective, this may influence the lung- and heart function of these patients.[66] In

the case of volcano ashes, short-term exposure can cause nose-, throat-, eye-, and skin

irritation, whereas long-term exposure might cause podoconiosis and sarcoma.[66-69] For

professional drivers, exposure to air pollution results in an increased risk of a heart

attack.[70-72] Even long-term exposure to smoke from cooking can cause severe health

effect due to particle inhalation, [73] and cigarette smoking is widely known to be toxic and

in many cases lead to lung cancer, genetic alterations, and asthma.[74]

It is well known that certain metals including for example Cu, Co, zinc (Zn), magnesium (Mg),

sodium (Na), potassium (K), calcium (Ca) and iron (Fe) are essential for humans. However,

when the dose of these elements becomes lower or higher than the a certain concentration

window, they may influence the health of humans, or growth of plants.[75] Aspects of

essentiality need to be considered, even though NPs made of these metals may need to be

considered as potentially more hazardous compared to their bulk materials due to the

15

specific properties of NPs as discussed above (e.g. high surface area).[12] Non-essential

widely used metals can in many cases be hazardous. Beryllium (Be) alloys used for electrical

parts and molds for plastics have as an example been shown to cause lung damage and

allergic reactions.[76] Lead (Pb) that exists in certain batteries, food and industrial emissions

is known to cause disability and kidney disease, [75] and exposure to Co, used in e.g.

batteries, car studs, hard metals and electronics, may cause asthma, acute illness and

interstitial pneumonitis.[76, 77] Cadmium (Cd) used in batteries, pigments and plastics can

cause lung irritation and liver damage,[75] and aluminum (Al) that is widely used in different

applications have been connected to both Parkinson dementia and Alzheimer’s disease.[78]

Nickel (Ni) and chromium (Cr) used for many applications may cause cancer.[75]

Cell uptake of NPs depends, as previously discussed on the size, shape, and composition of

the NPs.[79] Even the nervous system can take up NPs. Their origin is not only from

inhalation but also other from exposure pathways, like dermal penetration.[12] It has been

reported that inhaled metal NPs smaller than 30 nm can rapidly pass into the circulatory

system.[55, 80-83] NPs can also be transferred to different organs, such as the liver and

kidney.[84] There are further reports of NPs observed in the gastrointestinal tract and

skin.[68, 80, 85-88] NPs of Ag, TiO2, ZnO and Mg-oxides have been shown to have

antimicrobial activity.[89-91]

NPs are not always dangerous to human beings. Some positive effects of NPs have also been

observed and reported (non-toxic effects are seldom reported). Fullerene derivatives and

NPs made of substances containing oxygen vacancies can protect the neuro system and have

anti-apoptotic activity.[92, 93] Functionalized fullerenes have been shown to react with

oxygen species that attack lipids, proteins, and DNA, conferring neuroprotective

properties.[92, 93] It is important to stress that many NPs are non-toxic and exposure setting

dependent e.g. gold NPs or carbon nanotubes used for drug delivery.[94-96]

The toxic potency of NPs depends not only on the chemical environment but largely also on

the chemical composition, shape, size and particle ageing (weathering) to mention a few

factors.[12] Some of these factors are discussed below.

Which parameters determine the hazard of NPs? As discussed above, some NPs are toxic at certain conditions. However, it is inaccurate to

claim that any NP is toxic without considering for example the dose, the surface composition

as well as the particle size and shape. It is hence necessary to investigate which parameters

that affect the toxic potency of the NPs and for what conditions and exposure scenarios

these relations are valid.

The properties of the NPs

The toxicity of a substance is a product of its potency and dose. The dose is defined as the

amount of substance, and potency is the toxicity per amount of substance. The particle size

is largely governing the toxicity since smaller NPs typically induces more inflammation than

larger sized particles of the same material, partly due to a larger surface area.[80, 97-102] As

mentioned above, a special property of NPs are their high fraction of surface molecules. The

16

total surface area is therefore in general a better dose unit to use in nanotoxicity compared

with for example the total mass.[80, 97]

Agglomeration due to surface interactions of metal NPs has been mentioned above.

Agglomeration can increase the particle size and decrease the specific surface area, which

can affect the adsorption of e.g. biomolecules and the dissolution. Since the adsorbed

molecules and released metal ions/complexes are factors that influence the biouptake,

agglomeration can indirectly to some extent affect the toxic potency of metal NPs. [80, 82]

The chemical composition of the NPs and of the surface oxide, the interface towards the

environment) are properties that influence toxicity since they for instance influence the cell

uptake, chemical reactions with biomolecules, NPs localization, and particle ability (less

agglomeration and sedimentation).[79] The crystalline structure is also an important

factor.[98] Rutile and anatase are allotropes of TiO2. Rutile has been shown to induce DNA

damage in the presence of light, whereas no effects were observed for parallel exposures

with anatase. [98] The crystalline structure can form some NPs change after interaction with

aqueous compartments such as in the case of zinc sulfide (ZnS) NPs.[103] Since the surface

composition of metal alloys in most cases are very different from the bulk composition, toxic

properties of the pure metal constituents cannot be used to assess the properties of the

alloy.[104]

Metal release from the metal NPs surface

Metals will be released (dissolved) to different extent from any metal-containing surface,

including metal NPs, exposed to an aqueous adlayer or immersed in solution. It is hence

necessary to consider the extent (and metal speciation) of metal release when discussing the

toxicity of NPs. As an example, the toxicity Ag NPs is reported to, at least to some extent, be

related to the concentration of released Ag ions in the solution.[6]

The biouptake of metal ions and metal NPs by cells occurs in different ways. Usually, the

biouptake of metal NPs is affected by the extent and nature of particle agglomeration and

sedimentation unless there is a layer of adsorbed biomolecules or other ligands able to

stabilize the NPs in solution. The biouptake of metal ions is in contrast shown to be more

direct (faster diffusion and higher affinity to biological interface).[48] To assess particle

and/or metal release specific, these aspects need to be considered when assessing any

toxicity of metal NPs.[48]

Information on the metal NPs used in this PhD-project Stainless steel welding fume particles: Stainless steels are corrosion resistant Fe-based alloys

typically containing different amounts of mainly Cr, Ni, Mn, and Mo.[105] Stainless steel is

biocompatible and used in for example biomedical implants. Depending on grade, stainless

steel have superior corrosion properties due to its passive surface oxide with very good

barrier properties, and as a result these materials show a low extent of metal release and no

release of Cr(VI).[106] Stainless steel particles produced by inert gas or water atomized show

similar results.[107] Welding of stainless steel results in Mn- and Cr-rich fume particles of a

17

composition that can induce several respiratory diseases such as bronchitis, siderosis,

asthma, and possibly lung cancer.[108] This is due to its ability to induce oxidative stress and

inflammation due to the release of Cr(VI), Ni and Mn.[81, 109, 110] Fume particles formed

during welding of stainless steel are more toxic and reactive compared with fumes formed

from welding of mild steel.[108] Therefore, fume particles generated via welding of stainless

steels were investigated determine the chemical speciation of the surface oxide and of

released (dissolved) Cr in phosphate buffered saline.

Co and Co oxide NPs: Co metal is used in high wear-resistant alloys and as a binder in hard

metals because of its superior wear resistance, magnetic, and catalytic properties. Co NPs

and Co oxide NPs are used in pigments, catalysts, magnetic fluids, and as contrast agents for

medical imaging.[111] People that work in hard metal industries are at risk to be exposed to

airborne Co particles and may suffer from negative health effect.[112] Repeated exposure to

Co ions concentration exceeding 20 µg/L have been shown to cause risk for systemic

toxicity,[113] and exposure to Co can induce asthma and acute illness.[12] Tires with studs

made of WC-Co are used in Northern countries during the winter months to reduce

accidents on slippery roads. As a result of wear of the tire studs at the traffic settings,

nanosized particles of W and Co have been observed in road dust.[114-117] However, since

investigations of relevance for environmental and health risk assessments of Co NPs are rare,

one aim of this PhD-project was to provide novel knowledge on transformations of Co NPs at

different environmental settings. These studies were conducted on commercially available

powders of Co and Co oxide (Co3O4) NPs.

WC and WC-Co NPs: WC and WC-Co were commonly known as hard metal where Co acted

as binder metal. Hard metal had various kinds of applications such as drills and cutting

tools.[118] WC-Co was the material of tire studs used in winter.[119] Inhalation of WC-Co

particles could course lung diseases for hard metal workers.[120, 121] It had been reported

that WC-Co particles were more toxic than WC or Co due to generation of reactive oxygen

species[122] and a higher surface reactivity[123]. In this PhD thesis, WC-Co and WC particles

were used to investigate the adsorption of 2,3-DHBA and 3,4-DHBA.

Cu NPs: Cu is widely used metal. Copper-based nanostructured materials can be used in

conductive films, lubrication, nanofluids, catalysis, and also as potent microbicidal

agents.[124-126] Cu NPs had been found more toxic to zebrafish compared with the

corresponding amount of ionic Cu.[127] CuO NPs were more toxic than larger sized CuO

particles to the freshwater crustaceans Daphnia magna and Thamnocephalus platyurus.[128]

In this PhD thesis, Cu NPs was used to investigated the adsorption of 2,3-DHBA and 3,4-

DHBA as a comparison to WC and WC-Co NPs.

Co SCS NPs: The Co SCS were synthesized by one-step modification of the solution

combustion synthesis (SCS) approach, using hexamethylenetetramine (C6H12N4, HMT) as an

organic fuel/reducer and cobalt nitrate hexahydrate (Co(NO3)2·6H2O) (Co SCS sample) as

metal source/oxidizer. Briefly, in a typical experiment, 4.95 g of Co(NO3)2·6H2O was dissolved

in a minimum volume of hot distilled water and mixed with 1.39 g of HMT under constant

stirring. The reducer-to-oxidizer ratio was equal to 1.75 for all samples. The solution was fast

dried at 120°C until gel and then foam has formed. The dried foam in a heat-resistant beaker

18

was placed into preheated to 600°C muffle furnace in air atmosphere. After ignition, the

foam combust in an explosion self-propagate mode reaction resulting in light gray powder.

The powder was taken out from muffle and collected in a closed beaker to prevent metal

oxidation.[129, 130] The Co SCS NPs were for comparison with the Co NPs to study how the

carbon coated surface of the SCS NPs influences the transformation of NPs at FW conditions.

19

3. Experiments and Techniques

The transformation, fate and trophic transfer of the metal NPs listed in the previous section

were investigated in synthetic solutions or relevance for different environmental settings.

The composition and exposure conditions of all studied solutions are summarized in Table 1.

Table 1. Composition of synthetic solutions used in the studies and exposure conditions for

metal release experiments.

Name PBS Saline Simulated gut fluid

Tap water

Freshwater

Salts 8.77 g/L NaCl 1.28 g/L Na2HPO4

1.36 g/L KH2PO4

8.77 g/L NaCl

2.0 g/L NaCl 0.0146 g/L HCl (25%)

50 mg/L Alkalinity, HCO3 26 mg/L Chloride, Cl

0.0065 g/L NaHCO3, 0.00058 g/L KCl, 0.0294 g/L CaCl2·2H2O, 0.0123 g/L MgSO4·7H2O

pH pH 7.4 pH 7.4 pH 4 and 7 pH 10.2 pH 6.2

Temperature 37℃ 37℃ 25℃ 25℃ 25℃

Organic molecules

14.6, 146 mg/L amino acids, mucin, lysozyme, polylysine, poly glutamic acid

14.6, 146 mg/L amino acids, mucin, lysozyme, polylysine, poly glutamic acid

N/A 0.64 mg/L (TOC) excreted biomolecules from Daphnia magna, 450 µg/L (µg chlorophyll a /l) algae (Scenedesmus sp.)

15.412 mg/L 2,3-DHBA (0.1 mM), 15.412 mg/L 3,4-DHBA (0.1 mM)

10 mg/L Suwannee river natural organic matter

Exposure time 1 h, 24 h and 168 h

1 h and 24 h

5 min, 1h, and 24 h

5 min, 1 h, and 24 h

90 min 1 h, 6 h, and 24 h

Papers I, III I II II IV V

The metal NPs were exposed to different biological macromolecules (with different

functional groups and size) to investigate interactions between organic matter and the NPs

and to assess how adsorption of different biomolecules influences the dissolution and

trophic transfer. The investigated biomolecules with the different functional groups and

properties are listed in Table 2.

20

Table 2. Information of the investigated biomolecules (the reported charge is related to the

pH value of PBS and saline shown above)

Biomolecule Functional group

Lysine Amine (-NH3+)

Glutamine Amide (-CONH2)

Glutamic acid Carboxylate (-COO-)

Cysteine Thiol (-SH)

Poly lysine Amine (-NH3+), MW: 30 000-70 000 g/mol

Poly glutamic acid Carboxylate (-COO-), MW: 15 000-50 000 g/mol

Lysozyme Small compact globular protein, MW: 14 100 g/mol (net positively charged)

Mucin High molecular weight glycoprotein, MW: 7 · 106 g/mol (net negatively charged)

Algae Phytoplankton algal cell, containing glycoprotein, polysaccharides (negatively charged)

Excreted biomolecules from Daphnia magna

Degraded algae constituents (e.g. proteins, fatty acids and polysaccharides) (negatively charged)

2,3-DHBA and 3,4-DHBA Hydroxyl (-OH) and carboxyl (-COOH), a small degradation product of NOM

NOM Collection of large macromolecular structures: amino-, hydroxyl-, ketone-, phenolic-, and carboxylic functional groups

This PhD-study has employed a systematic approach combining surface, corrosion and

solution chemistry with state-of-the-art surface analyses. A wide range of different analytical

techniques was used, listed in Table 3, and briefly described below.

Table 3. Summary of the analytical techniques and main information provided in the

different investigations (Papers). Explanations of the abbreviations are given in the text

below.

Analytical Technique Given information Papers

ATR-FTIR Functional groups of adsorbed species

I-V

GF-AAS Metal concentration in solution (ppb(v))

I-V

Flame-AAS Metal concentration in I, IV, V

21

solution (ppm(v))

XPS Chemical speciation and elemental composition of outermost surface

I-V

Electrochemistry Corrosion information and surface chemical speciation of particles

II, IV, V

TEM-EDS Size and morphology of NPs I, II, IV, V

SEM Particle size and morphology

II, IV

XRD Bulk crystal structure of NPs I, IV, V

PCCS Size distribution of NPs in solution

I, III

NTA Size distribution of NPs in solution

I, V

Zeta potential Apparent surface potential of the NPs in solution

I-III, V

BET Specific surface area I-III, V

Speciation modelling of released Co in solution Joint expert speciation system (JESS), version 8.3,[131] was used for chemical equilibrium

speciation calculations of Co in PBS and amino acid solutions. The calculations were

performed for a temperature of 37 °C, and a redox potential of 300 mV based on

measurements using an Inlab redox electrode (Mettler Toledo, Sweden).

Attenuated Total Reflection Fourier Transform Infrared Spectroscopy (ATR-FTIR) Infrared spectroscopy is based on the absorption of infrared light due to the excitation from

the ground vibrational energy level to a higher energy level.[132] A molecular vibration will

absorb light, be IR active, when the dipole moment changes during the vibration. The

absorption of light gives information on molecular structure and functional groups. A peak

will be displayed in the spectrum due to the absorption of IR light at the frequency of the

vibration.[133]

22

Figure 7. Schematic illustration of the ATR-FTIR employed on a particle film.

ATR-FTIR spectroscopy is based on the total internal reflection phenomenon at the boundary

between two media.[134, 135] When a light transports through a medium and encounters a

medium with a lower refractive index, it undergoes a total internal reflection for incident

angles greater than the critical angle. The critical angle is calculated by:

θ𝑐𝑟𝑖𝑡 = 𝑠𝑖𝑛 −1(𝑛2

𝑛1)

where n2 is the lower refractive index and n1 is the higher. Although the incident light is

totally internally reflected at the interface, an evanescent wave penetrates a small distance

(around 2 µm) into the medium with the lower refractive index. It propagates perpendicular

to the surface in the plane of incidence.[136] ATR-FTIR is a useful instrument to investigate

the NPs surface condition in situ. It can probe the surface adsorption on the NPs surface in

environmentally and biologically relevant media. The technique provides adsorbed

molecular information that includes conformational and structural changes of coordinating

ligands.[133] The ATR-FTIR measurements were performed using a Bruker Tensor 37 with a

Platinum ATR accessory (diamond ATR crystal).

As shown in Figure 7, a film of NPs was prepared on the ATR crystal. A NPs stock solution

was prepared firstly containing 25 mg NPs and 10 mL ethanol. Tip sonication (Branson

Sonifier 250) was used to disperse the NPs in the stock solution. The sonication settings

include sonication time, sonication mode and output level. The stock solution was added

dropwise onto the ATR crystal by using a pipette with a total volume of around 300 µL.

Complete evaporation of ethanol was accomplished after 2 h. A flow cell was used to

introduce the solution of interest to the NPs film. A background spectrum of the NP film and

pure water was always collected and used for the ATR-FTIR measurements. The first step

was hence to introduce the background solution (ultrapure water) and collect the

background spectra with 512 scans at 4 cm-1 resolution, followed by introducing the sample

solution of interest collecting spectra with the same settings.

23

At the end of the measurements, the measuring cell was rinsed with a solution without

ligands to investigating irreversibility of adsorbed ligands onto the NP film. Rinsing was

performed using ultrapure water or background solution.

Atomic absorption spectroscopy (AAS) – dissolution / metal release Information of the total concentration of released metal into a given solution is important

when determining the dissolution and fate of NPs. Atomic absorption spectroscopy (AAS)

measures the concentration of metal ions in solution, both in sub-ppb (µg/L) and ppm (mg/L)

concentrations. The released (dissolved) metal concentration is divided by the total amount

of metal of the particles to give the released amount:

Released amount of metal =𝑑𝑖𝑠𝑠𝑜𝑙𝑣𝑒𝑑 𝑚𝑒𝑡𝑎𝑙 𝑐𝑜𝑛𝑐𝑒𝑛𝑡𝑟𝑎𝑡𝑖𝑜𝑛 𝑖𝑛 𝑡ℎ𝑒 𝑠𝑎𝑚𝑝𝑙𝑒

𝑡𝑜𝑡𝑎𝑙 𝑚𝑒𝑡𝑎𝑙 𝑐𝑜𝑛𝑐𝑒𝑛𝑡𝑟𝑎𝑡𝑖𝑜𝑛 𝑜𝑓 𝑡ℎ𝑒 𝑑𝑜𝑠𝑒 𝑠𝑎𝑚𝑝𝑙𝑒

The principle of AAS is that atoms absorb light at specific wavelengths.[137] Atomization of

samples is an important step in AAS measurement, and there are two atomization methods:

graphite furnace (GF) and flame, as shown in Figure 8. The selection of the atomization

method depends on the concentration of measured metal. After the atomization of the

samples, the quantity of the studied element is measured by absorption of light from the

atomized elements following Beer’s Law.[137, 138] A calibration curve is determined for the

quantification. Standard solutions of different metal concentrations are prepared, for

instance, the Co GF-AAS standard solutions were 10, 30, 60 and 100 µg/L Co in 1% ultrapure

HNO3. The LOD (the mean value of blank samples + three times the standard deviation of the

blank samples) is solution dependent and was for the systems investigated 0.5-1 µg/L for GF-

AAS, and 0.1-0.2 mg/L flame AAS. The AAS analysis was performed on a Perkin Elmer,

AAnalyst 800 instrument.

24

Figure 8. Schematic illustration of the GF-AAS and Flame AAS techniques to determine metal

concentrations in solution.

Exposure of nanoparticles in solution for metal release determinations Figure 9 depicts the different steps in sample preparation for assessing the extent and

kinetics of metal release from NPs by means of AAS. Tip sonication was used to disperse NPs

in the stock solution, using the same settings as for the preparation of stock solution for the

ATR-FTIR measurements. After the NPs dispersion, a specific volume of the stock solution

was transferred to the sample solution by using a pipette. One crucial step is to ensure that

the dose samples are prepared before transferring the stock solution to each group of

samples. The dose samples were used to quantify the amount of the transferred NPs by the

pipette since there can be a loss of NPs during the transfer due to poor dispersion, which

leads to a concentration gradient of NPs in the stock solution.[139] After the NPs were

transferred from the stock solution, the samples were exposed to the different solutions of

interest (Table 1). The NPs in the dose samples had to be digested, and the digestion

methods are described in each paper. Digestion for Co NPs was as an example performed

using 3.5 mL 65% ultrapure HNO3, 2 mL H2O2 and ultrapure water with a total volume of 10

mL. Aqua Regia (King’s water) was used for the SCS NPs and the welding fume particles.

After exposure, an ultra-centrifuge was used (Beckman Optima L-90K Ultracentrifuge, 50000

rpm, 1 h) to remove non-dissolved particles from the solutions. Finally, the supernatant was

collected as metal release samples for AAS measurements. Prior to AAS analysis, all

supernatants were acidified to a pH lower than 2.

25

Figure 9. Schematic illustration of sample preparation for the metal release experiments.

X-ray Photoelectron Spectroscopy (XPS) XPS was used to provide information on speciation and elemental composition of the surface

oxide of the NPs (ca. 5-10 nm information depth).[140] XPS is a technique based on the

photoelectric effect. An X-ray beam is introduced to the surface of the samples. The core

electrons of atoms at the surface absorb energy from the X-ray photons and if the energy is

large enough, the core electron will escape from its atomic energy level and be ejected.[141]

The emitted energy is measured by an electron spectrometer and the binding energy of the

electrons determined from the energy of the emitted electrons.

All XPS measurements were run on a Kratos AXIS UlraDLD instrument (Kratos Analytical) with

an Al Kα radiation source, and an X-ray photon energy of 1486.6 eV. All binding energies

were corrected to the adventitious carbon C 1s peak (C-C, C-H) set at 285.0 eV. Details for

the measurements are described in paper (I-V).

Electrochemistry – Open Circuit Potential (OCP) and Cyclic Voltammetry (CV) OCP is the potential of the working electrode (NPs) relative to a reference electrode when

there is no current flowing between the two electrodes.[142] OCP can provide information

on metal corrosion in a specific solution. For example, a lower OCP value of the NPs in a

system generally means that they are more easily corroded. OCP can help to understand the

NP dissolution and provide information on the effect of biomolecule adsorption onto the

NPs surface. A special electrochemical cell was used for measuring the OCP on the NPs,

Figure 10. The electrochemical setup consists of three electrodes, a working electrode, a

reference electrode, and a counter electrode. The working electrode consisted of metal NPs

attached on top of a paraffin-impregnated graphite electrode (PIGE), Ag/AgCl (sat. KCl) was

26

used as a reference electrode, and a platinum wire was used as the counter electrode. For

the preparation of the working electrode, the PIGE was first polished with 500 grit SiC paper

and then cleaned with paper tissues and ultrapure water. The PIGE was heated with a lighter

to soften the paraffin, after which the softened tip was pressed onto the NPs (typically 5 mg

NPs) to prepare the NPs film on the PIGE. An μAUTOLAB-TYPE III potentiostat was used for

the CV measurements of welding fume particles produced during stainless steel welding in

paper IV. Measurements for OCP were performed using a PARSTAT MC Multichannel

Potentiostat (Solartron Analytical) equipped with the VersaStudio software in paper II and V.

Figure 10. Schematic drawing of the electrochemical cell for electrochemical investigations of

NPs.

CV is a widely used electrochemical analytical method. It mainly investigates redox

transitions. During the measurement, a cyclic potential is applied to the working electrode.

Reduction takes place when the potential is scanned in the negative direction and oxidation

when the potential is scanned in the positive direction. The current is measured during the

scanning of the potential, creating a voltammogram. The reduction and oxidation reactions

are always accompanied by an electron transfer. The electron transfer in the reaction gives

rise to a current, which shows up as a peak in the voltammogram.[143]

In Paper IV, CV measurements were used to determine the chemical speciation of the

welding fume particle surfaces. NaOH (8 M, pH 13) was used as an electrolyte. The potential

was swept from OCP to -1.4 V (reduction), followed by oxidation to +0.2 V (in some other

cases the sample was first oxidized and then reduced). The step potential was set to 0.00244

V and the scan rate to 0.0005 V/s.

Transmission Electron Microscope - Energy Dispersive Spectrometer (TEM-EDS) TEM was used to investigate the morphology and size of the NPs. The principle of TEM is

more complex than optical microscopy. The optical microscope uses conventional lenses,

27

whereas the TEM uses electrostatic lenses. The electron beam is transmitted through a very

thin sample or substance on a grid. During the transmission, the electron interacts with the

atoms in the sample, causing scattering and absorption which leads to contrast in the

detected electron density in the obtained image. The contrast is related to the thickness and

the density of the sample, the atomic number, and the crystal structure or orientation.[144]

EDS is an X-ray analytical technique that provides quantitative information on the chemical

composition of a sample for the different atomic elements. An atom contains a ground state

in different energy levels surrounding the atom nucleus. Due to the electron beam

bombardment, an atom can be excited, which cause an electron from the inner energy level

to be ejected. Furthermore, the ejected electron leaves an electron-hole. An electron from a

higher energy level will occupy the hole, and the energy difference due to the movement of

the electron is released as an X-ray. The X-ray can be used to determine the elemental and

chemical composition of the material.[145]

A TEM microscope (Hitachi HT7700) operating at 100 kV was used for some of the

investigations (Papers I, V). A JEOL 200 kV 2100F field emission microscope operated in

scanning beam mode (STEM) combined with energy dispersive spectroscopy (EDS)

microanalysis were also used in this PhD project (Paper I).

Scanning Electron Microscopy (SEM) SEM generates images by rastering a focused beam of electrons across the sample. The

electrons interact with the atoms in the sample and produce elastic and inelastic scattered

electrons. Due to the interaction, several types of signals are obtained, including secondary

electrons (SE), backscattered electrons (BSE), Auger electrons, and X-rays. The SEM image is

based on these signals. For the surface characterization performed in this thesis, signals from

BSE were mainly used. BSE is formed due to elastic interactions, and it comes from the

deeper part of the sample (in general several micro meter deep information). BSE images

show high sensitivity to atomic number: the higher the atomic number, the brighter the

material appears in the image. It can also provide information about crystallography,

topography, and the magnetic field of the sample.[146]

The samples in this PhD project were mainly investigated using a tabletop scanning electron

microscope (SEM) with backscattered electron analysis (Hitachi TM-1000).

X-ray diffraction (XRD) X-ray diffraction can provide information about bulk crystal structure using the principle of

Bragg’s law:

𝑛𝜆 = 2𝑑sin𝜃

where n is a positive integer, λ is the wavelength of the incident wave, ϴ is incident angle

and d is the spacing between diffracting planes. When the X-ray is scattered from the crystal

lattice, peaks of the scattered intensity are shown when the incidence angle is equal to

scattering angle of a crystal plane, and the path length difference equals the integer number

28

of the wavelengths.[147] An X’Pert Pro PANalytical instrument was used for XRD

measurements in this PhD-project paper I, IV and V.

Photon Cross Correlation Spectroscopy (PCCS) PCCS (Nanophox, Sympatec GmbH, Germany) was used to measure the hydrodynamic size of

the particles in solution. The PCCS principle is based on dynamic light scattering (DLS), which

is based on Brownian motion. Brownian motion is the random motion of particles in a

solution or gas due to collision with moving molecules in the solution.[148] The instrument

estimates the diffusion coefficient of particles from the measured the light scattering. The

particle size can be obtained according to the Stokes-Einstein equation:

𝐷 =𝑘𝐵𝑇

6𝜋𝜂𝑟

where D is diffusion constant, kB is Boltzmann's constant, T is the absolute temperature, η is

the dynamic viscosity, and r is the radius of the spherical particle. PCCS measurements were

conducted in Papers I and III.

Nanoparticle Tracking Analysis (NTA) NTA (NanoSight NS300, Malevern, UK) was used to measure the hydrodynamic size of the

particles in solution. The principle of NTA is that it directly measures the speed (diffusion) of

the particles in solution by collecting a video of light scattering from solution. In this case

only the viscosity of the solution and the temperature influence, in addition to particle size,

the rate of the particle movement. Hence, the particle size is obtained directly instead of

getting a correlation function of the scattered light as in DLS. Moreover, it is not necessary to

use the Mie scattering theory to get the particle size distribution based on the number of

particles. NTA measurements were conducted in Papers I.

Zeta potential The zeta potential of NPs was determined by means of laser Doppler microelectrophoresis

using a Zetasizer Nano ZS instrument (Malvern Instruments, UK) at 25°C. The Zeta potential

can be used to gain information related to the surface charge. It needs to be noticed that the

zeta potential is measured at the slipping plane within the diffuse double layer of ions

outside the surface, and it is thus not the absolute surface potential. The Zeta potential is

measured based on the movement of charged NPs in an applied electric field. The

movement direction and velocity depend on particle charge, the solution, and the electric

field strength. The velocity is measured and can be used to calculate the zeta potential. The

Smoluchowski equation was used to calculate the potential from the measured

electrophoretic mobility of the particles. However, the Smoluchowski equation has its

limitations. It should be used for a thin double layer, i.e. the Debye screening length of the

29

diffuse double layer is smaller than particle radius, and it ignores surface conductivity.[149]

Zeta potential measurements were conducted in Papers I-III and V.

Brunauer–Emmett–Teller (BET) – surface area measurement BET measurement was used to analyze the specific surface area of NPs. BET is based on gas

multilayer physical adsorption of nitrogen molecules onto the NP surface.[150] BET

measurements were conducted in Papers I-III and V.

30

4. Key Results and Discussion 4.1. Characterization of particles before exposure. (Papers I-V)

The investigated NPs were prior to exposure into the different synthetic settings

characterized in terms of surface oxide, specific surface area, and particle size.

Compositional analyses of the surface oxide were acquired with XRD and XPS. These findings

are compiled in Table 4 together with the bulk composition of the NPs. The BET surface area

was determined for some of the NPs (Table 4). The results showed highly agglomerated NPs

when imaged in their dry state by means of TEM (Figure 11). TEM imaging of the Co SCS NPs

revealed a carbon layer with the thickness between 1-5 nm.

Table 4. Characteristics of the studied particles.

Name Surface speciation

Core speciation

Primary particle size (nm)

BET surface area (m2/g)

Papers

Welding fume particles

A complex oxide containing Cr, Mn, Fe etc.

N/A 10-1000 N/A IV

Co NPs Co3O4 Co 10-50 10.7 I-III

Co SCS NPs CoO or Co(OH)2

Co 50-500 1 V

Co3O4 NPs Co3O4 Co3O4 40-60 1.7 V

WC NPs WC, WO3, WO2

WC 40-80 1.76 II

WC-Co NPs WC, WO3, WO2, W-Co-O compound

WC-Co <100 2.67 II

Cu NPs CuO Cu <100 6.78 II

31

Figure 11. TEM images for some of the NPs investigated including welding fume particles, Co

NPs, WC NPs, WC-Co NPs, Cu NPs, Co3O4, and Co SCS NPs. A zoomed figure was shown for Co

SCS NPs to see the carbon shell. The scale bar is shown in each image.

The Co NPs agglomerated at ambient room conditions/manufacturing processes due to

strong van der Waals forces between similar particles. Hence, sonication was employed to

disperse the agglomerated particles in the solution and thereby obtain a dispersion of

particles with a smaller size, closer to their primary size (i.e. nanoscale particles). It is

important to elucidate if the sonication procedure influences the particle characteristics, for

example, changes the surface oxide characteristics and/or composition.

ATR-FTIR experiments were performed to determine the properties of the surface oxide on

the Co NPs before and after the sonication. As shown in Figure 12, the Co NPs revealed

bands in the ATR-FTIR spectra at 650-680 cm-1 and 550-600 cm-1, which are related to

Co3O4.[151, 152] The peak at 675 cm-1 was assigned to CoO.[151] The band at 620 cm-1,

related to Co3O4 [151], was visible prior to the sonication step but disappeared after the

sonication process. This may indicate a change of the Co3O4 (spinel) structure in the surface

oxide on the Co NPs. It was also observed that sonication further influences the extent of

dissolution of the Co NPs. Approximately 8 wt.% of the Co NPs was dissolved during the

sonication step when supplying an acoustic energy of 7056 J.[34] Although the sonication

step influenced the surface oxide composition and the dissolution of the Co NPs, the step

was necessary in order to study the particles on the nanoscale.

32

800 700 600 500 400 300

Co3O

4

Co NPs before sonication

Absorb

ance

Wavenumber (cm-1)

Co NPs sonicated in MQ

CoO

Figure 12. ATR-FTIR spectra for non-sonicated and sonicated Co NPs in MQ water.

33

4.2. Fate and transformation of Co NPs under simulated physiological conditions:

influence of biocorona formation (Paper I).

Co NPs is a reactive material which may be harmful to human health (Co ion concentrations

exceeding 20 µg/L are reported to result in risks for systemic toxicity.[7]) Therefore, it is

important to elucidate any health and environmental risks induced by Co NPs.

Understanding of the effect of biocorona formation on the stabilization and dissolution of

the Co NPs is important to investigate to assess the fate and transformation of Co NPs in

biological and environmental media. Environmental systems contain different processes,

which are rarely at equilibrium. Changes in metal speciation are often dynamic processes

since complexation/dissociation readily take place in solution in parallel with the

adsorption/desorption of the biomolecules. The dynamics and kinetics of changes in surface

characteristics of NPs are hence key factors to investigate.[133] PBS is a widely used

synthetic fluid used to simulate the physiological conditions of human blood (The phosphate

concentration in PBS was 20 mM which is higher than at more realistic conditions, ca. 0.38

mM [153]), and the addition of biomolecules can, at least to some extent, imitate a more

realistic situation since the adsorption of biomolecules will affect the dynamics taking place

at the NP surface.[135, 154] An improved knowledge of interactions taking place between

different biomolecules and the Co NPs in different synthetic fluids can provide important

information for further studies, such as trophic transfer, uptake or toxicity of Co NPs

(schematically illustrated in Figure 15).

The adsorption of biomolecules onto the Co NPs was studied by employing ATR-FTIR. Figure

13a shows the spectra of Co NPs in PBS solution. Peaks originating from adsorbed phosphate

are marked with a box in the figure. Vibrational bands at around 1111 cm-1 and 957 cm-1

were observed, which correspond to stretching vibrations of the P-O bonds.[155] The peaks

shifted around 30-40 cm-1 relative to the position of the P-O bonds in solution, which

indicate the formation of Co-phosphate surface complexes.[156, 157]

34

Figure 13. ATR-IR spectra of a) Co NPs exposed (2 h) to PBS and PBS containing 146 mg/L

lysine. b) Co NPs exposed (2 min) to saline (0.15 M NaCl) and MQ water containing 146 mg/L

lysine. c) Co NPs exposed (2 min) in PBS containing 146 mg/L polylysine, polyglutamic acid,

lysozyme or mucin

The interactions between the Co NPs and the amino acids with different functional groups

were studied to deduce how the different characteristics of the amino acids influence the

adsorption in PBS (Paper I). Lysine, with a positively charged amine functional group, was

expected to adsorb onto the negatively charged Co NPs, due to an electrostatic attraction

between the ionic groups. However, the ATR-FTIR spectra of Co NPs in PBS solution in the

presence of lysine did not show any detectable vibrational bands corresponding the amino

acid (Figure 13a). This indicates a strong complexation of phosphate with the Co NP surface,

which prevents the adsorption of lysine. In order to understand the effect of phosphate ions

in the solution, measurements were performed on Co NPs with lysine in MQ water and in

saline (0.15 M NaCl) solution (pH 7.4) (Figure 13 b). Clear vibrational bands corresponding to

lysine were observed both in saline and in MQ water. Evidently, the formation of strong Co-

phosphate surface complexes hinders the small amino acids from adsorbing onto the Co NPs.

The same effect was observed for the other investigated amino acids, i.e. no or minor

adsorption in PBS, but evident adsorption in MQ water (Paper I).

In contrast to the amino acids, adsorption of polypeptides and protein onto the Co NP film

was observed in PBS (Figure 13c). This was concluded based on the presence of bands

35

related to Amide I and Amide II, marked in the spectra.[7] Vibrational bands originating from

phosphate could still be observed in the spectra from which follow that the interaction

taking place between the Co NPs and phosphate existed as well. The phosphate peak

position in the case of polylysine was slightly different, which could be related to the

coordination between phosphate and the positively charged amine group in the polylysine

side chain. This did nevertheless not induce any change in the vibrational mode of HPO42-,

similar to observations when it coordinates with the Co NP surface, but resembled rather

uncoordinated HPO42-.

The difference in adsorption between amino acids and polypeptides/proteins show that the

larger biomolecules have a higher affinity for the Co NP surface. For the

polypeptides/proteins this may be explained that they have more possible contacts points

with the surface and that the driving force for adsorption is mainly due to an entropy gain

from the release of small molecules such as water and counter ions from the interface

between the molecules and the surface.[158] However, other surface interactions such as

ionic (electrostatic) interactions (both repulsive and attractive), hydrogen bonding,

hydrophobic interactions, hydration forces, acid-base interactions, and van der Waals forces

cannot be excluded.[39] It was furthermore observed that the negatively charged

polyglutamic acid and mucin adsorbed onto the negatively charged Co NPs, indicating that a

gain in entropy dominates the interaction during adsorption.

The amount of released Co ions in solution was determined by means of AAS to assess if the

adsorption of phosphate and the formation of a biocorona (adsorption of biomolecules) on

the Co NPs affected the dissolution of the Co NPs. Figure 14a shows the released amount of

Co ions in saline and in PBS, respectively. The results clearly show that the released amount

of Co was higher in the presence of phosphate, which indicates the existence of interactions

between Co NPs and phosphate. One explanation for this might be that adsorbed phosphate

can destabilize the surface oxide of the NPs, which in turns leads to an increased release of

Co ions due to weakening of the Co – O bonds. In addition, the formation of Co-phosphate

complexes can enhance the dissolution since the system is pushed further away from the

equilibrium in terms of Co solubility. Around 45% (per mass) of the Co NPs was dissolved in

PBS solution already within 1 h. The dissolution was almost complete after 1 h since the

released amount only increased with another 10% after 24 h exposure in PBS.

36

Figure 14. Amount of released Co: a) 10 mg/L Co NPs in PBS and saline (pH 7.4) for

1 h and 24 h. b) 10 mg/L Co NPs in PBS and in PBS containing 146 mg/L lysine, glutamine,

glutamic acid or cysteine (pH 7.4) for 1 h and 24 h. c) 10 mg/L Co NPs in PBS and in PBS

containing 146 mg/L polylysine, polyglutamic acid, lysozyme or mucin (pH 7.4) for 1 h and 24

h.

In the presence of amino acids (Figure 14b), the released amount of Co was close to that in

pure PBS except for cysteine. The dissolution showed a similar tendency, rapid dissolution

during the first hour of exposure followed by slower rates during the next coming 24 h. The

dissolution of the Co NPs exposed to cysteine was very high, more than 85% in PBS. The

other amino acids had less effect on the dissolution of Co NPs in PBS.

To investigate the formation of Co-phosphate complexes dominating the surface interaction

in the presence of amino acids, JESS chemical equilibrium calculations (pH 7.4, 37 °C) were

performed to assess the chemical speciation of Co(II) ions (0.1, 1, 10 mg/L) in PBS and in PBS

containing the different amino acids (146 mg/L) (Paper I). The results showed that Co ions

have a higher affinity to phosphate compared to the functional groups of the amino acids,

except for cysteine (the thiol group). However, the ATR-FTIR measurements did not show

any adsorption of cysteine. This might be due to that the calculation was related to Co ions,

while Co oxide is the main constituent on the Co NP surface that might have different

properties. However, the increased release of Co in the case of cysteine can correlate with

37

the higher tendency for complexation between Co and the thiol group in cysteine, forming

soluble complexes. Hence, this will push the saturation of dissolution further away from

equilibrium, from which follows that more Co ions can be released.

These findings show that it is not only the Co oxide on the surface of the NPs but also the

formation of Co-phosphate surface complexes that need to be considered for Co NPs in

solutions containing phosphate. Amino acids could affect the Co NPs dissolution only if the

specific amino acid has a higher affinity than phosphate to Co ions by pushing the dissolution

balance away from equilibrium. Otherwise, the interaction between phosphate and Co NPs

will dominate the system. Moreover, Co NPs dissolve rapidly in PBS, which should be

considered in, for instance, health risk assessments. Since more than 40% of Co NPs were

dissolved after 1 h in the solution, the toxic potency of the Co NPs may originate from both

Co NPs, Co ions and/or Co-phosphate complexes.

Figure 14c shows the dissolution of Co NPs in PBS solution containing polypeptides, lysozyme

and mucin. Compared to pure PBS, the dissolution was reduced during the first hour of

exposure. As shown above, the results show that both the polypeptides and the proteins

adsorb on the Co NPs (Figure 13c). Thus, the adsorption initially hinders, at least to some

extent, the dissolution of the Co NPs. However, for longer exposure time periods (24 h), the

released amount of Co was similar to parallel measurements in pure PBS.

The data imply that the negatively charged amino acids and polypeptides, glutamic acid and

polyglutamic acid, resulted in somewhat higher released quantities after 1 h compared with

the positively charged lysine and polylysine in Figure 14 b and c. These findings are in line

with previous findings that show the metal−surface interactions of negatively charged

proteins (BSA) to significantly enhance the extent of metal release from both stainless steel

(of different grades) and silver, whereas the interaction with positively charged proteins

(lysozyme) predominantly resulted in a minor increase of the amount of released metals

from stainless steel and a significant reduction of released silver from massive silver.[159-

161]

In summary it is concluded that polypeptides/proteins can affect the interactions between

Co NPs and phosphate, by adsorbing onto their surfaces, as opposed to amino acids (Figure

15). Initially, this will due to changes in surface characteristics lead to a reduced extent of

dissolution of the Co NPs. Changes in surface characteristics and reactivity will most likely

influence both the toxic potency and/ or the biouptake of the Co NPs. Furthermore, for

health risk assessment it is very important to consider effects of large biomolecules (e.g.

proteins) and not only effects of inorganic salt constituents in the human body flood.

Understanding of the interaction between NPs and biomolecules may be a key factor that

decides the surface characteristics and fate of the NPs.

38

Figure 15. Schematic illustration of the interaction between Co NPs and biomolecules.

39

4.3. Fate and transformation of metal NPs in simulated environmental conditions:

influence of ecocorona formation (Paper II and unpublished results)

Nanoparticles of relevance in traffic settings

Metal NPs can readily interact with natural organic matter (NOM) present in aquatic settings

such as freshwater (FW). The aim of this study was to investigate the fate and

transformation of metal NPs of relevance for traffic settings (Paper II). Engineered NPs

relevant for NPs originating from studded tires, tungsten carbide cobalt (WC-Co), was

compared with NPs of tungsten carbide (WC), cobalt (Co) and copper (Cu). The adsorption of

dihydroxy benzoic acid (DHBA), a small degradation product of natural organic matter, onto

different metal NPs was monitored in-situ with ATR-FTIR in FW. Adsorption of NOM on the

NPs may significantly affect their mobility and bioavailability in the environment.

Figure 16. ATR-FTIR spectra of Co NPs (a), WC-Co NPs (b), WC NPs (c), and Cu NPs (d) in

SW+2,3-DHBA and SW+3,4-DHBA after 90 min immersion.

The spectra in Figure 16 show that DHBA adsorbed on the WC-Co NPs, Co NPs and Cu NPs,

with peak positions appearing at 1240-1260 cm-1 (phenol C-O(H) stretch), 1365-1380 cm-1

(symmetric COO--stretch), 1410-1440 cm-1 (symmetric COO--stretch), and 1490-1515 cm-1 (C-

C ring stretching vibration/asymmetric COO-). For the WC NPs, only peaks originating from

WO3 were observed between 810 to 950 cm-1. It was also observed that DHBA rapidly

40

adsorbed onto the Cu NPs and the Co NPs. On the WC-Co NPs, however, the adsorption was

significantly lower, and minor adsorption was observed within 90 min. In contrast, the

adsorption became higher after 24 h (known from zeta potential results in paper II) due to

changes in surface characteristics (surface oxide). No adsorption of DHBA was observed to

take place onto the WC NPs.

AAS measurements of the dissolution of Co from the WC-Co NPs showed that Co was

completely released within a few hours. All investigated NPs released more than 1 wt-% of

their total metal content, and the amount of released metal increased in the order: WC < Cu

< Co < WC-Co NPs. It was also shown that the release of Co in the presence of DHBA was

higher compared to exposure in FW only. This may be explained by a weakening of the

surface oxide through ligand exchange due to adsorbed DHBA. A similar effect was observed

for the Cu NPs, whereas the Cu NPs were more dependent on the presence of DHBA (Paper

II).

Core-shell, oxide, and composite Co nanoparticles behave differently in freshwater in terms of

dissolution and NOM adsorption.

Core-shell Co NPs, Co NPs prepared by solution combustion synthesis (SCS), and Co oxide

NPs were investigated in order to understand how NPs surface properties and composition

influence their behavior in solution. The behavior of the NPs was studied in FW conditions

with and without the addition of NOM (Suwannee River NOM).

The core-shell NPs consisted of a surface oxide covering a metal core, as illustrated in Figure

11 and Table 4. For the SCS Co NPs, the particles were embedded in a layer of amorphous

carbon that was formed during the synthesis. NPs of the Co oxide consisted of Co3O4. The

surface oxide on the core-shell NPs was mainly Co3O4, but had a different structure

compared to pure oxide (Co3O4 NPs, Table 4).

Figure 17 shows spectra from the ATR-FTIR investigations of the three different NPs in FW and in FW with NOM. Adsorption of carbonate and sulfate could be observed in FW for all NPs.

41

Figure 17. ATR-FTIR spectra of Co NPs, Co SCS NPs, and Co3O4 NPs in a) pure freshwater and

b) freshwater with 10 mg/L NOM. The pH value of the solutions was adjusted to 6.2. The

spectra were collected during 3 h after the solution was introduced onto NPs film. The

background spectra were for all cases ultrapure water with films of the respective NPs.

The band between 1100 to 1140 cm-1 corresponds to the symmetric stretch of sulfate.[162] This vibrational band was observed for all NPs in FW.

The symmetric CO32- stretching vibration is located in the range between 1045 and 1080 cm-1

(sometimes overlapping with sulfate at ca. 1100-1140 cm-1), while the asymmetric carbonate

stretching is split into several peaks due to interaction between carbonate and the surface,

resulting in a change in symmetry of the CO32-.[163, 164] The degree of split of the

asymmetric stretching band gives information on the adsorption geometry of carbonate. The

split for Co3O4 and the SCS Co NPs was approximately 100 cm-1, which is indicative of a

monodentate surface coordination of carbonate.[163] The Co NPs had a larger split (ca. 170

cm-1) which could indicate a slightly different coordination geometry of adsorbed

carbonate.[163]

Rinsing with ultrapure water at the end of the experiment (6 h), resulted in mostly

unchanged spectra even though the sulfate band (1100-1140 cm-1) was slightly reduced in

absorbance.

The addition of NOM to FW changed the spectra in Figure 17, with the major changes being

a reduced absorbance of the bands at 1050-1140 cm-1 (less adsorbed carbonate and sulfate)

and the addition of bands related to NOM (mostly carboxylates). There was some overlap

between the carboxylate stretches from NOM and asymmetric carbonate stretches in the

range between 1350 and 1600 cm-1. NOM was nonetheless clearly seen to adsorb based on

the asymmetric COO- stretch observed at approximately 1547-1595 cm-1,[165] as this band

was absent in FW, with the exception of Co NPs in FW where the asymmetric carbonate

stretch was observed in this spectral region as well. However, the lack of symmetric

carbonate stretch for the Co NPs in FW containing NOM (ca. 1060 cm-1) shows that the band

42

at 1595 cm-1 originates from asymmetrical stretching mode of carboxylate of NOM. NOM is a

heterogeneous mixture of different molecules, mainly composed of fulvic and humic acids,

rich in carboxylate groups.[165] Since rinsing with MQ water had little effect on any of the

bands, the adsorption of NOM was concluded to be relatively strong.

A band occurred at 1469 cm-1 for the Co3O4 NPs, at 1481 cm-1 for the Co SCS NPs, and at

1479 cm-1 for the Co NPs. This band position is relatively close to asymmetric carbonate

stretching vibrations (Table 2). However it cannot be excluded that this band also contains a

contribution from an asymmetric vibration of carboxylate, which indicates an additional

coordination mode (bridging bidentate or bidentate[165]) of the carboxylate group to the

surface.[5] This tentative assignment of carboxylate was especially valid for the Co NPs that

did not show any evidence of carbonate adsorption in the presence of NOM.

From the results it was concluded that the adsorption of NOM dominated in FW, see Figure

18b, findings in agreement with MINTEQ calculation results (unpublished results) that

showed a higher affinity of Co ions to NOM compared to both carbonate and sulfate. Zeta

potential measurements showed furthermore a more negative zeta potential for the three

types of Co NPs in the presence of NOM, which agrees with the observed adsorption of NOM

(Table 5).

Table 5. Zeta potential of the NPs in freshwater with and without the addition of NOM.

Particle Freshwater Freshwater + NOM

Co NPs -12.6 ± 3.7 -18.8 ± 1.0

Co SCS NPs -7.8 ± 1.0 -14.1 ± 1.5

Co3O4 NPs -12.8 ± 0.8 -24.6 ± 1.2

NOM adsorption did not only change the surface composition but also the particle size

distribution. NTA results (Table 6) collected after 6 and 24 h showed smaller particle sizes in

the presence of NOM, which may be due to electrostatic repulsion in addition to steric

stabilization of the NPs as a result of NOM adsorption. The results agreed to ZnO NPs[166]

and CuO NPs[167].

Table 6. Mean particle sizes (nm) in FW with and without the addition of NOM, exposure

time periods: 1 h, 6 h and 24 h.

Particle Freshwater 1 h

Freshwater + NOM 1 h

Freshwater 6 h

Freshwater + NOM 6 h

Freshwater 24 h

Freshwater + NOM 24 h

Co NPs 225 ± 58 139 ± 13 327 ± 68 111 ± 3 623 ± 84 91 ± 8

Co SCS 225 ± 58 160 ± 10 230 ± 51 146 ± 7 292 ± 64 191 ± 8

Co3O4 201 ± 57 151 ± 15 285 ± 79 143 ± 13 214 ± 51 140 ± 8

43

Figure 18 shows the dissolution of the different Co NPs in FW and FW with NOM. The Co3O4

NPs resulted in Co ion concentrations lower than the detection limit, which indicates very

limited dissolution. More Co was released after 1 h for the Co NPs compared with the Co SCS

NPs, but vice versa after longer exposure time periods (24 h), both in FW and in FW

containing NOM (p < 0.05).

1 h 24 h0

10

20

30

40

50

60

70

80

90

100

***

Co

dis

so

lutio

n (

% o

f to

tal C

o)

Co NPs in FW

Co NPs in FW + NOM

Co SCS NPs in FW

Co SCS NPs in FW + NOM

Co3O

4 NPs in FW

Co3O

4 NPs in FW + NOM

*

Figure 18. Amount of dissolved Co per loaded mass of Co NPs, Co SCS NPs and Co3O4 NPs in

freshwater (FW) and freshwater with NOM. Star (*) in the figure corresponds the AAS data

was below LOD.

OCP results of the different Co-based NPs are presented in Figure 19 (including the OCP of

the PIGE electrode without NPs as a reference). As previously described, a higher OCP value

generally correlates with improved protective properties of the surface oxide. In some cases

this could possibly mean a tendency of lower dissolution.

44

0 1 2 3 4 5 6 7 8-0.3

-0.2

-0.1

0.0

0.1

0.2

0.3

0.4

Co SCS NPs in FW

Co SCS NPs in FW + NOM

PIGE in FW + NOM

Co NPs in FW

Co NPs in FW + NOM

PIGE in FW

Co3O

4 NPs in FW + NOM

Pote

ntial (V

vs. A

g/A

gC

l sat. K

Cl)

Time (h)

Co3O

4 NPs in FW

Figure 19. OCP results for Co NPs, Co SCS NPs and Co3O4 NPs in FW (solid) and FW + NOM

(dashed line). The NPs were exposed in the solution for 6 h. The OCP of the PIGE electrode

without NPs is included as a reference.

The Co3O4 NP film showed the highest OCP, and nearly the same with and without NOM.

These findings comply with the results on dissolution results that showed no detectable

amounts of dissolution. Co SCS NP showed the lowest OCP value among the three different

NPs, but its dissolution was not lower than observed for the Co NPs (Figure 18). The addition

of NOM increased the OCP for both the Co NPs and the Co SCS NPs. Adsorption of NOM onto

the NPs could thus hinder diffusion of species in the electrochemical dissolution processes

(for example oxygen and metal ions) and as a consequence increase the OCP value. However,

as shown in Figure 19, the Co SCS NPs showed the lowest OCP value but did not release

significantly more Co compared to the Co NPs. The lack of correlation between OCP and

dissolution is similar to findings of Cu corrosion in the presence of NOM. It was seen that

NOM increased the OCP on the same time as the release of Cu was increased. This was

explained by the influence of NOM on the formation of surface oxide and decreased the

adsorption of carbonates, as well as solubilization of Cu by NOM.[168] This implies that

other processes (e.g. ligand-induced chemical processes) govern the dissolution rather than

electrochemical processes.[169]

45

The surface area available for dissolution processes will largely influence the dissolution rate

of the NPs.[170] The presence of NOM in FW decreased the size of the particle agglomerates

compared with FW only for all the NPs, which theoretically should result in an enhanced

release of Co. However, the presence of NOM did not increase the dissolution, which may

imply that the influence of surface area for the given conditions is small compared with

other mechanisms influenced by the addition of NOM to FW. Another possibility is that any

effects are hidden behind e.g. the increased OCP and the reduced surface area.

NOM can also influence the release process via chemical, non-Faradaic processes.[171]

Complexation through several contact points for the same molecule (e.g. humic acid) to the

surface, can in general result in reduced dissolution since all of the contact points need to be

detached at the same moment in order for a ligand exchange to take place. However, the

small sized molecules of NOM could possibly complex and promote ligand-induced

dissolution[172, 173] (see also previous section). Complexation of NOM to the surface of the

Co3O4 NPs (Figure 3) did not result in any detectable release of Co (Figure 5). If there would

be any enhanced release of Co from the Co3O4 NPs due to surface complexation with NOM,

it was not strong enough to cause any detectable change in the extent of released Co.

The dissolution of Co NPs and Co SCS NPs was not accelerated by addition of NOM, which is

opposite to observations for ZnO and Cu NPs.[51] However, it is important to notice that the

loading of NPs in many previous studies was high enough so that the solubility of metal was

reached. The addition of NOM for these high loadings will increase the dissolution due to

increase of the metal species solubility from the formation of soluble metal-NOM species.[51]

In summary, even though carbonate and sulfate adsorbed to all studied NPs in FW, the

adsorption of NOM dominated in all cases. The adsorption of NOM hindered the

agglomeration of the different Co-based NPs and contributed in all cases to a more negative

zeta potential. Compared with FW only, the adsorption of NOM resulted in an improved

corrosion resistance for both the Co NPs and the Co SCS NPs.

The composition of the surface oxide of the Co NPs (Co3O4, CoO) and the SCS Co NPs

(CoO/Co(OH)2+carbon layer) compared with Co3O4 NPs is suggested to be a very important

characteristic when it comes to environmental transformations of these NPs. The dissolution

was much higher for the NPs with the surface oxide composed of CoO and Co(OH)2

compared with Co3O4, as these oxides in general are less stable than Co3O4. This trend was

also in line with literature findings on the transformation (and toxicity) in synthetic biological

media in which Co NPs and CoO NPs dissolved faster and showed higher toxicity towards

lung cells compared with Co3O4.[174]

46

Figure 20. Schematic drawing of the behavior of Co NPs (core-shell), SCS Co NPs and Co3O4

NPs in FW and in FW+NOM.

47

4.4. Trophic transfer of Co NPs in the aquatic food web (Paper III) The trophic transfer is one aspect of the risk assessment of Co NPs in the aquatic food

web.[175] An aquatic food web containing algae (level 1), zooplankton (Daphnia magna)

(level 2) and fish (Crucian carp) (level 3) was selected in this work.[10, 176] The food web

transfer was divided into three steps, as shown in Figure 21 including i) heteroagglomeration

of Co NPs/ions with algae, ii) Co uptake by Daphnia magna, and iii) Co uptake by fish. Since

biomolecules can adsorb onto the Co NP surface and influence their transformation and

surface properties, as discussed in the previous section, excreted biomolecules from

Daphnia were added to the algae solution to investigate their influence on Co NP food web

transfer.

Figure 21. Schematic illustration of the transfer of Co NPs in the aquatic food web, and the

influence of the addition of excreted biomolecules.

Trophic level 1: algae

Studies on the first trophic level involved the interaction between Co NPs and algae

(Scenedesmus sp.), excreted biomolecules, and algae with excreted biomolecules in tap

water. Tap water was used in this study since it was the same water used for the uptake

studies in the animals. ATR-FTIR measurements were used to investigate the surface

interactions in these solutions (Figure 22).

48

Figure 22. ATR-FTIR spectra of Co NP films after 2 h exposure at pH 10.2 in a: tap water. b:

tap water with algae (450 µg/L chlorophyll a/L). The dashed line refers to conditions with

pure algae without Co NPs. c: tap water with the excreted biomolecule (0.64 mg/L TOC). The

dashed line reflect spectra of a bulk solution of excreted biomolecules (0.64 mg/L TOC),

without Co NPs. d: tap water containing both algae and excreted biomolecules.

Figure 22a shows the spectra of Co NPs in tap water at pH 10.2 after 2 h. The pH was fixed to

10.2 to match the pH, induced by the presence of algae in tap water, due to removal of

carbon dioxide during photosynthesis. The pH of lake water rich in algae is similar to this

value.[177] Carbonate adsorption was evident based on peaks observed at 1582 cm-1, 1421

cm-1 and 1380 cm-1[164]. The board peak at around 1080 cm-1 could have the contribution

from both carbonate and sulfate.[162] The adsorption peaks from algae (Figure 22b) were

observed at around 1683 cm-1,1556 cm-1, 1500 cm-1, 1384 cm-1 and 1043 cm-1 and related to

Amide I, Amide II, CH2 bending, Amide III, symmetric COO- and C–O–C, C-OH, C-O, C-C

stretching vibrations, respectively.[133, 178, 179]

The spectra of Co NPs in tap water in the presence of excreted biomolecules are shown in

Figure 22c with main vibrational bands at 1675 cm-1, 1568 cm-1, 1385 cm-1, 1032 cm-1 and

1080 cm-1. These peak positions are similar to findings for the Co NPs with algae, which

indicate that molecules of similar functional groups (e.g. in proteins, polysaccharides) are

adsorbed in both solutions.

Spectra of Co NPs in tap water containing both algae and excreted biomolecules are

presented in Figure 22d. The marked peaks relate to carbonate adsorption from the tap

49

water, algae adsorption, and adsorption of the excreted biomolecule. As mentioned above,

there are overlaps between vibrational bands originating from the different solutions.

However, it can nevertheless be concluded that the algae and excreted biomolecule

dominate the adsorption, as seen from the peak positions and the relative peak intensities.

When comparing the spectra in Figures 22b-d, the peak shapes at around 1040 cm-1 are

similar in b and d. The relative peak intensity is moreover similar also for the peaks at around

1570 cm-1, 1385 cm-1 and 1040 cm-1, indicative of a predominance of algae adsorbed onto

the Co NP film. A more detailed discussion is given in Paper III. In all, algae adsorbs to the Co

NPs in tap water, which may influence the heteroagglomeration of Co NPs with algae.

Adsorption onto the Co NPs was also evident for the excreted biomolecules.

Zeta potential measurements showed similar values of the apparent surface potential in tap

water with and without excreted biomolecules, Table 7. These results, together with the

ATR-FTIR measurements with and without excreted biomolecules, suggest a minor, if any,

influence of the excreted biomolecules on the surface properties of the Co NPs in the

presence of algae.

Table 7. Zeta potential of Co NPs in tap water containing algae solution with and without

excreted biomolecules.

5 min 1 h 24 h

Co NPs + algae -17 ± 2 mV -14 ± 2 mV -13 ± 3 mV

Co NPs + algae + excreted biomolecules

-15 ± 2 mV -14 ± 1 mV -14 ± 2 mV

Co can exist in several different forms if Co NPs are dispersed into an aquatic setting, e.g. as

individual-, agglomerated- or sedimented NPs, as Co ions released from the NPs or as labile

or strongly formed complexes. Interactions between Co NPs/ions/complexes and algae

should be considered as well since these processes potentially result in heteroagglomeration

between the Co NPs and the algae.[180, 181] These different forms of Co and Co NPs are

schematically illustrated in Figure 23. The main interest for trophic transfer is Co NPs/ions

agglomerated with algae, as algae serve as food for Daphnia magna (the next trophic level).

50

Figure 23. Schematic illustration of the results from the partition and transformation of Co

NPs (6.2 mg/L) with algae in tap water after 24 h.

The extent of Co heteroagglomeration with algae was compared to the fate of Co NPs in

pure tap water. It was found that the fraction of Co agglomerated with algae was only 0.4-5%

of the total added Co (Paper III). The addition excreted biomolecules did not affect the

interaction. The results furthermore showed that more than 80% of the Co NPs had settled

from solution (sedimented) within 24 h of exposure in solution (algae and tap water). This

could partly be one reason to why the extent of heteroagglomeration of Co NPs with algae

was low.

AAS measurements showed that less than 1 wt-% of the NPs was released as soluble Co ionic

species after 24 h. This could be related to a low dissolution of Co from the Co NPs or that

released Co ions were bound to the algae and removed from solution by the centrifugation

process. When comparing the amount of released Co in tap water containing algae with, for

example, to the dissolution of Co NPs in PBS (Paper I) it was evident that the amount of

released Co ions was significantly lower. One explanation was the higher pH value for the tap

water containing algae solution (10.2), which most probably affected the release of Co.

An adsorption of biomolecules on the NPs can influence their ability to heteroragglomerate

with algae. For example, Au NPs coated with positively charged polyallylamine hydrochloride

(PAH) showed very high heteroagglomeration with living algae cells due to electrostatic

attraction.[52] Besides the electrostatic interaction, the hydrophobic effect can influence

these interactions. The hydrophobic part of humic acid (HA) is reported to show a high

affinity to the algae cell walls and lead to a higher agglomeration of Au NPs coated by HA.[52]

However, as shown in Paper III, the NOM adsorption onto the Co NPs did not increase the

heteroagglomeration with algae. However, opposite of findings for the Co NPs that readily

sedimented, the Au NPs was stabilized by the coating and did not sediment from solution.

51

Trophic level 2: Daphnia magna

In the next studied trophic level, Daphnia magna was fed with algae that had been exposed

to Co NPs for 24 h (Figure 24). The Co uptake is presented as the total concentration of Co

and therefore contains a contribution from both Co NPs and Co ions/complexes.

Figure 24. The uptake of Co by Daphnia, given as g Co per g Daphnia dry body weight. Inset:

enlarged figure showing the control samples. There were 20 Daphnia magna in each sample,

and 8 samples were prepared for every group (control, TW+algae, TW+algae+excreted

biomolecules). The control Daphnia magna were fed with pure algae. The other Daphnia

were fed by algae mixed with Co NPs in tap water, with and without the presence of excreted

biomolecules. All Daphnia were exposed for 24 h. The error bars reflect values of one

standard deviation. The asterisks indicate statistically significant differences (p<0.05,

Student’s t-test).

The control samples of Daphnia that were fed by pure algae showed an uptake of 2·10-7 g

Co/ g dry weight, which was around 300 times less than samples with Daphnia feeding on

algae that had been exposed to Co NPs. The Student’s t-test showed that the addition of

excreted biomolecules did not result in any statistical difference in Co uptake by Daphnia.

A previous investigation following a similar exposure protocol as for the Co NPs showed an

approximately 20 times higher uptake for WC NPs compared with the Co NPs investigated in

this study.[182] The uptake in Daphnia of Au NPs with different surface stabilizing coatings

were 8.3·10-3 g Au/ g dry weight (citrate), 10.5·10-3 g Au/ g dry weight (HA), and 23.6·10-3 g

52

Au/ g dry weight (PAH) after 24 h feeding. This uptake was also significantly higher than

observed for the Co NPs (Figure 24). However, it was shown that after depuration of

Daphnia, the uptake of Au NPs decreased by a factor of at least five.[52] The results indicate

that Daphnia magna can remove uptaken particles and that the initial differences in surface

chemistry (different capping agents) do not affect the uptake of the Au NPs. [52]

The low uptake of Co NPs compared to the WC NPs and the Au NPs may be caused by the

different surface properties of the NPs and the sedimentation rate. It was shown that the Co

NPs exhibited a high sedimentation rate (Paper III). Hence, the difference in uptake could be

caused by a large amount of sedimentation of the Co NPs compared with the more

colloidally stable Au NPs.[52]

Trophic level 3: Crucian carp

Fish (Crucian carp) fed by the Daphnia containing Co (see the previous section), was the third

trophic level in this study. The Co uptake in the stomach and intestine of the fish is shown in

Figure 25, as these two organs showed this highest Co uptake. None of the other organs

investigated (i.e. gills, muscle, blood, blood plasma, and brain) showed any significant

increases of Co uptake compared with the control samples.

Figure 25. Co uptake in the fish stomach and intestine presented as g Co per g in dry body

organ weight. a: stomach, b: intestine. The fish were fed 8 times with Daphnia (which was

fed by algae exposed to Co NPs for 24 h) for 15 days. The error bars reflect the values of one

standard deviation. The asterisks indicate statistically significant differences (p<0.05,

Student’s t-test). The control samples were fed in the same way using “pure” Daphnia

(without Co NPs).

The amount of Co uptake in the control samples (fish fed with Daphnia without exposure to

Co NPs) was ca. 10-7 g Co/ g dry weight, an amount at or below the detection limit of the AAS

quantification of Co, and similar to the control samples of the Daphnia. No significant

differences in Co concentration were observed in the stomach compared to the control

samples. However, there was a statistically confirmed increase in the amount of Co in the

53

intestine compared with the control samples. The Co uptake in the intestine and stomach

was not seen to be affected by the addition of excreted biomolecules (Paper III).

The bioaccumulation factor for the transfer between Daphnia and the Crucian carp was

calculated as the Co uptake by the fish (g Co/ g dry weight) divided by the uptake in

Daphnia.[56, 183] The Co uptake in Daphnia was approximately 100 times larger compared

to the Crucian carp per weight basis. Hence, there was no bioaccumulation of Co in the

trophic transfer from Daphnia to the Crucian carp. In the view of risk assessment, these

results imply a low risk for Co NPs accumulation at a high trophic level when considering fish

feeding on Daphnia. Despite the fact that the investigated dose of Co NPs (6.2 mg/L) in this

PhD-project was substantially higher compared to reality (nature sea water contains 0.025

ng/L Co and surface water contains 0.041 ng/L Co in Japan [184]; 0.4-9 µg/L Co was

determined in Bega river during 2016 [185]), the amount of Co uptake in Daphnia was only

in the order of 10-4 g Co/ g dry weight, which was 20 times lower compared to WC NPs in

previous studies.[182] Even in the fish stomach, the amount of Co was not statistically

different from observations of non-Co-exposed fish with typically 10-6 g Co/ g dry weight.

The dissolution of Co was investigated in simulated gastric fluid (SGF) to elucidate the

influence of pH of the fate of Co NPs that were taken up in the trophic levels 2 and 3 (Figure

26). A pH of 4 represents the conditions of fish intestine,[186] and a pH of 7 is similar to the

pH in the gut of Daphnia.

pH 4 pH 70

20

40

60

80

100

Co r

ele

ase (

% o

f to

tal C

o)

SGF 1 h

SGF + excreted biomolecules 1 h

SGF 24 h

SGF + excreted biomolecules 24 h

*

*

Figure 26. Amount of released Co after 1 and 24 h in simulated gastric fluid (SGF) with 2.004

g/L NaCl representing the gut of Daphnia (pH 7) and the stomach of Crucian carp (fish) (pH 4)

in the absence and presence of excreted biomolecules. The asterisks indicate statistically

significant differences (p<0.05, Student’s t-test). The Co NP loading was, on average, 6.2

mg/L for all the samples.

The results showed that approximately 60% of the Co NPs was dissolved after 1 h in SGF at

pH 4 (Figure 26). The dissolution was significantly lower at pH 7 compared with pH 4, both

with and without the presence of excreted biomolecule. At pH 7, the presence of excreted

54

biomolecules reduced the extent of dissolution. This may be a consequence of changes in

surface characteristics as a result of the adsorption of the biomolecule (see the previous

section) and that influences the dissolution of the Co NPs. The adsorption of the excreted

biomolecules (ATR-FTIR results at pH 10.2) may form a barrier on the Co NP surfaces that

hinders the dissolution process. Similar effects have been observed for Ag, for which the

adsorption of lysozyme onto massive silver resulted in a reduction of the released amount of

silver ions.[160] The barrier effect became less evident at pH 4. The reason could be related

to changed interactions between the Co NPs and the biomolecules at pH 4. The charge of the

excreted biomolecules may as an example be different, which can lead to a different

conformation of the adsorbed biomolecules on the Co NPs surface.

Results in SGF suggest that a majority of the Co NPs taken up by Daphnia could remain as

particles as only 10 % of the Co NPs dissolved within 1 h at pH 7 and 20 % after 24 h.

However, the Co uptake in Daphnia was not affected by the presence of excreted

biomolecules despite the fact that they may influence the dissolution of Co NPs in the

Daphnia. More than 60 % of the Co NPs in the fish stomach was dissolved within 1 h, which

indicates that the Co uptake in the fish stomach was a result of both Co NPs and Co

ions/complex. The Co release did not change after 24 h compared with 1 h, which means the

dissolution of Co NPs in the fish stomach was rapid and reached equilibrium conditions

within 1 h.

In short, no bioaccumulation of Co was observed in the studied aquatic food web, with a

large fraction of the Co NPs rapidly settled (sedimented) upon entry into the aquatic setting

(algae and tap water) and thus unavailable for any trophic transfer to Daphnia. Even though

the Co NPs loading was high (10 mg/L) the bioaccumulation from Daphnia to fish was low.

This accumulation was not affected by the sedimentation of the Co NPs since the

sedimented Co NPs could not be transferred to the Daphnia. The presence of excreted

biomolecules did not affect the Co transfer in the food web, although their adsorption onto

the Co NPs was evident. Surface interactions with algae, which were more abundant, were

probably more important than the excreted biomolecules.

60% of the Co NPs were dissolved during 24 h in the fish stomach as estimated from the

synthetic gastric fluid exposure. The corresponding dissolved amount of the Co NPs was only

10-20 % in the gut of the Daphnia. Hence, the capacity of dissolution in a trophic level is a

key factor to consider when assessing the biouptake and bioaccumulation of NPs.

The studied scenario used is substantially higher concentrations of Co NPs (10 mg/L) than

observed in any realistic scenario. Nonetheless, the lack of bioaccumulation of Co in fish

suggests a low risk of risks for this trophic level in the case of the dispersion of Co NPs to an

aquatic setting. Instead, other trophic levels or bottom-feeding organisms should be

considered for targeted risk assessments.

55

4.5. Dissolution of stainless steel welding fume particles in PBS solution (Paper IV) The surface characteristics of metal-containing NPs is a key factor that affects its behavior in

solution. Dissolution and surface characteristics were investigated for fume particles

incidentally generated during welding of stainless steel. Different size fractions of the

welding fume particles were collected as a function of different filler metals, base alloys and

shielding gases, Table 8.

Figure 27. Schematic illustration of metal release from welding fume particles into the lung

compartment.

Table 8. Welding parameters for the collected welding fume particles generated when

welding of stainless steel (test A, B and C)

Test Welding method

Base metal Filler metal Shielding gas

A MAG LDX2101 FCW-LDX2101 CORGON 18

B MAG 304L FCW-308L MISON 18

C MAG LDX2101 FCW-LDX2101 MISON 18 MAG: gas metal arc welding using an active shielding gas FCW: flux-cored wire CORGON 18: Ar + 18% CO2 MISON 18: Ar + 18% CO2 + 0.03% NO LDX 2101: A duplex stainless steel that consists of both austenite and ferrite. It was chosen as represent for duplex stainless steel which had a different resistance during welding, and higher Mn content. (21.6% Cr, 1.6% Ni, 4.9% Mn, 0.027% C) 304L: Austenitic stainless steel with high chromium and low carbon content. (18.3% Cr, 8.1% Ni, 1.1% Mn, 0.018% C) 308L: Austentic stainless steel with low carbon content.(19.2% Cr, 10.5% Ni, 1.1% Mn, 0.028% C)

56

CV was used to analyze the surface speciation of the welding fume particles. The

measurements started at the OCP, then proceeded in the negative direction (reduction) until

the hydrogen evolution started (-1.4 V vs. Ag/AgCl sat. KCl, the reference electrode referred

to hereafter), followed by scanning in the positive direction (oxidation) until oxygen

evolution started (0.2 V). A broad reduction peak at around -1.05 V was observed for tests A,

B and C. The oxidation peaks at -1.15 V and -0.95 V were related to the oxidation of

previously reduced iron oxide (during the reduction step).[187, 188] Observed oxidation

peaks at approximately -0.45 V and -0.2 V were assigned to oxidation of reduced manganese

oxide,[188, 189] and the narrow peak at -0.55 V to the oxidation peak of reduced bismuth

(hydro)oxide.[190] A completely oxidized surface oxide was evident based on CV

measurements scanned in the opposite direction (first oxidization and then reduction)

(Paper IV).

Figure 28. a: Cyclic voltammograms of collected welding fume particles (test A, B and C) in

8 M NaOH (pH 13) for the 400-640 nm size fraction of the welding fume particles when

reduced from their OCP followed by oxidation. b and c correspond to magnified

voltammograms.

The broad reduction peak shown in Figure 28b at –1.05 V was not related to a solitary metal

oxide since the oxidation peaks for Fe, Mn and Bi were observed, Figure 28c, and their

57

corresponding reduction peaks would have been at different positions. Therefore, the

surface oxide of the welding fume particles could be composed of a complex mixed oxide

containing at least Fe, Mn, and Bi. The results clearly showed, due to a missing reduction

peak at -0.75 V, the lack of any solitary hexavalent chromium oxide within surface oxide of

the welding fume particles.

PBS was used to simulate the human lung environment. It has similar ionic strength as

human blood and a pH of 7.4, which is also relevant for human lung conditions,[191] but

without any reducing and/or complexation capacity that would disable the detection of

Cr(VI). Measured amounts of released metals in PBS (Ni, Mn, Fe, Cr(VI) and Cr(III)) compared

with their total amounts in the collected welding fume particles are displayed in Figure 29.

0

5

10

15

20

25

30

35

40

45

50

Test C

#10

Test C

#4

Test B

#10

Test B

#4

Test A

#10

Meta

l re

lease

in P

BS

(wt%

of to

tal m

eta

l o

n f

ilte

r)

Ni

Mn

Fe

Cr(VI)

Cr(III)

Test A

#4

Figure 29. Released amounts of Fe, Cr(VI), Cr(III), Mn, and Ni in PBS (24 h, 37 °C, pH 7.4) for

the welding fume particles of different size fractions (#4 (60-108 nm) and #10 (1000-1600

nm)).

The welding fume particles released predominantly Cr and Mn. A higher amount of released

Ni was observed from the nano-sized particles compared to the micron-sized particles and

the agglomerates. No evident particle effect was observed in relation to the release of Cr,

Mn, and Fe compared with their corresponding metal contents (determined after complete

particle digestion). The release of Cr in PBS was dominated by trivalent and hexavalent

chromium (up to 70% of the total contained Cr in particles) followed by Mn (0 to 30%), Ni (0

to 20%), and Fe (0 to 3%) (Paper IV). The release of Cr(VI) from the welding fume particles

can result in adverse effects on human health.[192, 193] Later studies of similar welding

fume particles showed that cytotoxicity could be correlated to the release amount of Cr(VI)

58

in PBS.[194] The welding fume particles were shown to be cytotoxic and induce DNA damage

in human lung cells as well.[194]

Tests A, B, and C were characterized by very different welding settings in terms of base

metal, filler metal and shielding gas. As shown in Paper IV, the base metal and consequently

the filler metal composition evidently influence both the solubility of the welding fume

particles and their composition. Later studies have shown an evident influence of the arc

mode (melting rate) during welding, while the importance of the shielding gas was less

important from a solubility and compositional perspective.[194]

To conclude, fume particles formed during welding of stainless steel have a complex mixed

surface oxide composed of at least Mn, Fe, Cr, and Bi, but most probably also Si oxides

(Paper IV). The use of different welding settings (such as base metal, filler metal or shielding

gas) results in welding fume particles of different characteristics and dissolution properties.

Cr was predominantly released as Cr(VI) in PBS within 24 h. In contrast to the release of Mn,

Cr and Fe, the release of Ni was influences by the particle size.

59

Concluding remarks The main objective of this doctoral thesis was to provide an understanding of reactive metal

NPs interactions in biological and environmental solutions. Results obtained within the

framework of this thesis are expected to be used in risk assessments of engineered NPs.

The following main conclusions were drawn based on the results and discussions of this

thesis:

Phosphate used as a buffer in PBS can largely influence the dissolution and surface

properties of Co NPs. PBS enhanced the extent of dissolution and phosphate was adsorbed

to the surface of the Co NPs. The addition of amino acids did not affect the dissolution

properties or the surface composition due to strong interactions between Co and phosphate,

with the exception of cysteine. Larger sized biomolecules (polypeptides, proteins) became

adsorbed to the Co NP surfaces which initially reduced their dissolution in PBS. Adsorption of

the larger biomolecules took place due to mainly hydrophobic and ionic interactions,

combined with an entropy gain from the release of smaller molecules. It is thus important to

consider the influence of these larger biomolecules for any predictions of the environmental

fate of Co NP dispersion into aquatic settings.

Heteroagglomeration was evident to take place between the Co NPs and algae in tap water

in quantities of typically 0.4-5% of the Co NPs (6.2 mg/L loading) and sedimentation was

substantial. No bioaccumulation of Co NPs or its dissolved species were observed in the

trophic transfer from algae (Scenedesmus sp.), to zooplankton (Daphnia magna), to fish

(Crucian carp). The particle loading was however higher compared to real exposure

scenarios, conditions that favors homoagglomeration compared with heteroagglomeration

due to the high concentration of NPs. Nonetheless, the lack of bioaccumulation in fish

suggests that there is a low risk of adverse effects for this trophic level given a case of

dispersion of Co NPs to an aquatic setting. Instead, other trophic levels or bottom-feeding

organisms should be considered for targeted risk assessments.

The addition of excreted biomolecules did not affect the Co NPs transfer in the studied food

web. Although the adsorption of the excreted biomolecules was observed on the Co NPs, the

extent of agglomeration of the Co NPs with algae, biouptake of Co in Daphnia or fish organs

were not affected. Interactions between the Co NPs and the algae were considered to be of

higher importance. This was partly related to the relative low concentration of excreted

biomolecules (TOC equal to 0.64 mg/L).

NOM adsorbed onto the Co NPs, Co SCS NPs, and Co3O4 NPs. This adsorption was

predominant due to a higher affinity between NOM and Co compared to carbonate and

sulfate (constituents of FW). The adsorption of NOM reduced the extent of agglomeration of

the NPs, but had only a minor effect on the dissolution (reduced in the case of the Co NPs).

The surface oxide composition of the Co NPs (Co3O4, CoO) and SCS (CoO/Co(OH)2) compared

with Co3O4 NPs was shown to be a very important to assess any environmental

transformations and/or fate of these NPs. The dissolution was substantially higher for the

NPs with a surface oxide composed of CoO and Co(OH)2 compared to Co3O4, despite their

higher solubility.

60

The welding fume particles generated during welding of stainless steel had highly oxidized

surfaces consisting of complex oxides composed of Cr and Mn. Cr was predominantly

released from stainless steel welding fume particles in PBS solution as trivalent and

hexavalent Cr followed by Mn, Ni and Fe. The nano-sized welding fume particles released

higher amounts of Ni compared to the micron-sized particles. No evident effect on particle

size could be observed in relation to the release of Cr or Mn.

61

Future work In this thesis, the behavior of reactive metal NPs in terms of dissolution, biomolecule/NOM

adsorption, and agglomeration was investigated in biological and environmental fluids.

Future aspects that need to be taken into account are summarized below:

The behavior of Co NPs TW, FW and PBS solutions was investigated in this PhD-project.

However, it was difficult to compare the data between the different solutions, since the

pH of these solutions were different, which may be a key factor. In future work it would

be interesting to control the pH value and compare the behavior of the Co NPs in

different fluids. This could be done in TW, FW and PBS solutions of pH 4, 6, 7 and 9. Such

approach may more clearly show the influence of pH on the behavior of the Co NPs, and

in parallel investigate the behavior of the Co NPs in the different media.

In Paper I, the dissolution of Co NPs in PBS solution was investigated, and a rapid

dissolution was observed. For the ATR-FTIR experiments, a negative peak at around 1650

cm-1 was observed that was a result of the release of the Co NP film on the ATR crystal. It

would be interesting to investigate if the amount of released Co could be connected to

the ATR-FTIR spectra. One idea was to collect the solution that was introduced into the

ATR flow cell and to use AAS to analysis the Co concentration in the collected solution to

determine the release of Co at different time points. The area of the negative peak could

then be calculated and be related to the Co NPs dissolution in order to find a statistic

relation between the data.

The behavior of the Co NPs was investigated at laboratory conditions in this PhD-project.

It should be interesting to move one step forward to obtain data at more realistic

conditions. For instance, interactions between the Co NPs and microorganisms such as

bacteria would be interesting to study. In Paper III, trophic transfer of Co NPs was

studied. Since microorganisms are an important trophic level in nature, investigations of

the interaction between Co NPs and microorganisms would be helpful for risk

assessments of Co NPs. Furthermore, agglomeration was observed for the Co NPs in the

aquatic setting and a large amount of Co NPs rapidly sediment from the solution. Thus, in

order to investigate trophic transfer of Co NPs, bottom-feeding organisms should be

considered.

For NPs in simulated biological media, the exposure conditions might need some more

modification. For instance, linear rocking conditions do not really simulate human blood

conditions. There will always be a flow in the human blood stream, which means that

effects of movement should be considered as well.

For the dissolution assessment of the welding fume particles (from welding of stainless

steel), it is necessary to add biomolecules into the PBS solution in order to take the

presence of biomolecules into account when simulating the human lung. In Papers, I, II

and V, effects of biomolecules/NOM adsorption were investigated, and it was clearly

observed that the presence of biomolecules cannot be ignored in order to understand

the fate and behavior of the NPs.

Many electrochemical experiments were employed during this PhD study in order to

provide supporting information to the NP behavior assess using different surface

analytical techniques. However, the results were not very satisfactory due to the lack of

62

reproducibility. Many different factors including NP dissolution, NP corrosion,

biomolecules/NOM adsorption on the NPs, and even biomolecules/NOM adsorption on

the working electrode, and biomolecules/NOM oxidation in the system, had to be

considered in the measurements that made it difficult to interpret the results. This was

especially evident for the OCP experiments that were not stable or reproducible. Thus

these electrochemical measurements, at least the OCP experiments, may not be suitable

to study the behavior of metal NPs. However, some modified settings are worth trying.

For instance, the NPs could be pre-exposed to biomolecules in solution prior to the

electrochemical testing. Centrifugation could then be used to remove the particles from

the solution. The NPs could be collected, and as long as the adsorption of

biomolecules/NOM is strong enough, there should be an adsorbed layer on the surface

of the NPs. The NPs with adsorbed molecules could then be attached to the PIGE

electrode, and the electrolyte in the electrochemical cell should be absent of

biomolecules/NOM. In this way, the influence of biomolecule/NOM adsorption on the

working electrode and of biomolecule/NOM oxidation is avoided. Hence, dynamic

interactions taking place within the system should be a bit simpler.

In order to get some general information on the tendency of the behavior of the metal

NPs in biological and environmental fluids, more types of NPs and biomolecules should

be investigated. The collection of a database is possible, but will be very ambitious. It

would though be very useful in risk assessments. Even though it will be a huge amount of

work, it should be valuable since the results are of relevance for a sustainable and

environmental-friendly development of the society, and of importance for improved

human health.

63

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